Abstract
Previous studies have demonstrated positive net primary production
effects with increased precipitation in semi-arid grasslands of Inner
Mongolian. The knowledge of the store and storage potential of carbon
(C) and nitrogen (N) can help us better understand how ecosystems would
respond to anthropogenic disturbances under different management
strategies. Therefore, we carried out research on the storage of organic
C and N in four sites where the floras and landform were similar but the
intensities of disturbance by grazing animals varied. The primary
objective of this study was to pinpoint how the store and storage
potential of C and N would respond to grazing exclusion and
precipitation. So, the concentrations of soil organic carbon (SOC) and
soil total nitrogen (STN) were determined in the 0–50 cm soil layers,
and the concentrations of microbial biomass carbon (MBC) and microbial
biomass nitrogen (MBN) were measured by an innovative method in our
study. Additionally, the contents of soil bacteria and fungi were
determined in the 0–50-cm soil layers.
The research showed that the total C, N, MBC and MBN storage were
significantly different among the four grasslands
(P <0.05), and they all decreased substantially with
grassland degradation and increased to a significant extent with the
introduction of natural grassland (NG). More than 90% of C and 95% of
N stored in soil were lost, while they were minor in other pools
(including those stored in the above-ground biomass, litter, and roots).
It is interesting to note that micro-aggregate was a limiting factor to
soil and microbial nutrients pool compared to precipitation. The limit
range of C and N storage observed in these grassland soils suggested
that enclosed-fence might be a valuable mechanism of sequestering C in
the top meter of the soil profile. The results of this study can provide
a basis for better recovery of the grasslands that grazing disturbed in
semi-arid areas.
Keywords : soil organic carbon; microbial activity; semi-arid
steppe
1. Introduction
About 40% of the land surface in the world are covered by grasslands
(White et al., 2000), most of which are in drylands (Maestre et
al.,2012) and sustain the main livestock-production systems (Kemp et
al., 2013); Soil contains more carbon than the atmosphere and vegetation
combined (Averill C et al., 2014). In some areas, grasslands may serve
as important global C sinks. For those in the tropics, the annual sink
is about 0.5 Pg of C and it is basically influenced by the baseline of
SOC level and annual precipitation (Davidson et al., 1995; Scurlock and
Hall, 1998; Tan et al., 2006). Moreover, they store 200–300 Pg of soil
carbon, which influences the global carbon cycle significantly (Scurlock
and Hall, 1998).
The significant impacts of the differences in land-use and ecosystem
management strategies on the storage of C in grasslands has been shown
clearly in the past few decades (Lugo and Brown, 1993; Post and Kwon,
2000; Jones and Donnelly, 2004; Billings, 2006; Elmore and Asner, 2006;
Liao et al., 2006). Meanwhile, nitrogen is vital for vegetation
productivity as well as terrestrial ecosystem stability (Harpole et al.,
2007). Due to the close connection of soil C and N cycles, there is
substantive concern that the changes in land-use and its integrative
effect may alter the storage of C and N in the soil (Houghton et al.,
1999). For instance, in the northern China, the severe degradation or
desertification of temperate grasslands was just led by rapid livestock
expansion (Li, 1994; Dong and Zhang, 2005). The improvement of soil
nutrient availability can help the plant recover from a disturbance,
promote the fast-growing plants to regrow faster given a improved soil
water status, and then facilitate the recovery of plant growth in an
ecosystem (Stampfli et al., 2018). Obviously, nitrogen and water
availability is of important impacts on the net primary productivity
(NPP) of grasslands (Xu et al., 2014), especially those in semi-arid
regions where annual precipitation inputs are significantly less than
evaporation (Heisler-White et al., 2008). After the drought, N
availability could be triggered; meanwhile, there would be improvements
in plant absorption, reallocation and N-use efficiency (Mackie et al.,
2019). The soil N availability could affect the composition of microbial
community and the richness of soil bacteria and fungi.
However, there is a dearth of information regarding the potential of C
and N storage due to the absence of stable or mature grassland
ecosystems. The knowledge of the store and storage potential of carbon
(C) and nitrogen (N) can help us understand how ecosystems would respond
to anthropogenic disturbances under different management strategies.
Therefore, we conducted a study on organic C and N storage in four sites
that were floristically and topographically similar to ascertain the
impact of grazing exclusion and precipitation on the store and storage
potential of C and N. This study can provide a basis for recovery of the
grasslands that grazing disturbed in semi-arid areas.
2. Materials and Methods
2.1 Study Site and Experiment Design
The study site (42°02′27″N, 116°17′59″E, elevation 1,334 m a.s.l) is
located in the south of Duolun County of Inner Mongolia, northern China.
The site has a semi-arid climate with mean annual precipitation of 379.4
mm and mean annual temperature of 2.1 °C, ranging from −17.8 °C in
January to 18.8 °C in July. The mean annual precipitation from January
to June is 118.8 mm and January to August 306.7 mm. The mean annual
temperature from January to June is 0.1 °C and January to August 4.69
°C. The soil pH of the experimental site is ranging from
7.1(P =0.04) to 7.4(P =0.003) under precipitation. It is
featured by a chestnut soil (Chinese Soil Taxonomy) or Calcisorthic
Aridisol (the U.S. Soil
Taxonomy). The steppe in that region
has been severely degraded due to overgrazing in the past 50 years. In
this study, four experimental sites were selected and subjected to SD,
MD, LD, and NG respectively (Table 1). Site SD had been exposed to
long-term heavy grazing; and an estimated 90% of the above-ground
biomass had been consumed by livestock every year. As indicated, the
grasslands were severely degraded by the extremely sparse vegetation
coverage (<10%). Site MD had also been exposed to long-term
heavy grazing, with an estimated 75% of the above-ground biomass
consumed by livestock every year. It was moderately degraded with an
existing vegetation cover of 10–25%. Site LD had been subjected to
long-term free-grazing, and an estimated 65% of the above-ground
biomass had been consumed by livestock every year with existing
vegetation cover of 25–30%. Influenced by climatic conditions and
human activities, the dominant vegetation on aeolian soil was Spiraea
saliclfolia and Salix gordejevii. Associated species were Leymus
chinensis, Agropyron cristatum, etc. Site NG was set up in 2000 by
fencing a 40 ha of a previously free-grazing grassland when the local
government initiated a grassland protection program (Xu et al. 2012).
Natural grassland (NG) is dominated
by needlegrass (Stipa krylovii), wheatgrass (Agropyron cristatum), and
prairie sagewort (Artemisia frigida). It is undeniable that there are
pseudo-replication issues given that there is only one plot per grazing
regime, and in these studies, this problem is quite common. However, it
should be certain that changes in SOC and STN among the four plots in
this study are mainly caused by grazing intensity and length of
exclusion because the four experimental plots are floristically and
topographically similar and all are distributed in the same upper basalt
platform (Table 1)
2.2 Field Sampling and Laboratory Analysis
In early April 2017, we selected representative plots at site SD, MD,
LD, and NG to measure the above-ground and below-ground C and N contents
in plants, litter and roots. The field samplings conducted in mid-June
and mid-August 2017 in Inner Mongolia were taken as the research
objects. In each plot, 5 sampling quadrats (each 1 m * 1 m) were set up
at 10-m intervals along a random transect. (The information on
vegetation and soil types can be acquired fromhttp://www.maplet.org .) The above-ground samples in plants and
litter were collected subsequently. Root samples were determined by a
soil corer (diameter 7 cm), with 5 sampling points for each site.
Similarly, soil sampling was conducted by a soil sampler (diameter, 4
cm), and the soil samples were separately collected from five layers at
the depths of 0–10 cm, 10–20 cm, 20–30 cm, 30–40 cm, and 40–50 cm
in each sampling point.
2.3 Chemical Analysis
(1) The pH of 0–10 cm soil samples, in H2O (soil :
water 1:5), was tested with a PHS-3S pH meter (Sartorius, Germany).
(2) Soil aggregates were fractionated by a laser particle size analyzer.
For particle-size fraction (i.e. into sand, silt, and clay), 50 g of
soil (<2mm) was dispersed in 250 ml of distilled water with a
KS-600 probe-type Ultrasonic Cell Disrupter System (Shanghai Precision
& Scientific Instrument Co. Ltd., China) set at 360 W; then the
different particle-size fractions were detached as per Morra et
al.(1991). After isolation, large macro-aggregates
(>2000μm), small macro-aggregates (250–2000μm), and
micro-aggregates (<250μm) were extracted.
(3) The contents (%) of organic C in the samples of plant, litter,
root, and soil were measured by a modified Mebius method separately
(Nelson and Sommers, 1982). Then, 0.5 g samples were digested with 5mL
SOC solution. The concentration of SOC was determined by chemical
oxidation with K2Cr2O7solution. Then, approximately 0.2 g air-dried soil was weighed in
separate test tubes, with analytical replication for each sample and a
total of 5 sub-samples from each treatment. Exactly 10 mL of reaction
solution containing 0.032 Mol Ag2SO4,
0.06667 Mol K2Cr2O7, and
9.39 Mol H2SO4 was added to each tube
which was then placed in a hot (~200 °C)
H3PO4 bath for 5 mins. Three tubes
containing 0.5 g SiO2 were also used as blanks. The
amount of K2Cr2O7consumed by SOC oxidation was determined by titrating the remaining
K2Cr2O7 in the test
tubes after digestion. The SOC oxidation efficiency was determined to be
92%, thus a 1.08 correction factor was used. The total N (%) of plant,
litter, root, and soil was measured with the modified Kjeldahl wet
digestion procedure (Gallaher et al., 1976) with a 2300 Kjeltec Analyzer
Unit (FOSS, Sweden). Then, 0.5 g of air-dried and finely ground soil, in
duplicate, were taken for each sample. Afterwards, the soil samples were
digested with 18.76 Mol H2SO4 for 6 hrs
under progressively elevated temperature from 150 °C (1 hr), 270 °C (1
hr) to 380 °C (4 hrs) and then automatically distilled in a Kjeldahl
apparatus where the evolved NH3 was adsorbed by
H3BO3 (20 g L−1). The
yield of NH3 was titrated by diluted
H2SO4 (0.02 Mol) and converted into the
total amount of nitrogen in soil.
(4) Soil microbial biomass analysis. Soil microbial biomass C (MBC) and
N (MBN) were determined by the fumigation-extraction method (Vance et
al. 1987). In this study, an innovative method, a soil microbial
fumigation device and an improved fumigation method (ZL 2020 3
0785811.3) were employed to extract microbial biomass. During the
process of microbial fumigation, the traditional glass drying dish needs
to be sealed with paraffin wax, which is not only poor in sealing, but
also difficult to control the temperature of constant temperature
culture. Therefore, vacuum pump is required in the process of culture to
maintain a certain negative pressure. However, it would be challenging
to maintain the same batch of samples under the same negative pressure,
thus resulting in greater individual differences in the fumigation of
samples. Additionally, the result of removing chloroform from the
container is not ideal after fumigation. It could be claimed that the
employment of the patented device could definitely resolve the defects
of the prior art. Then, 10 g of each soil sample was fumigated with
ethanol-free chloroform (CHCl3) for 24 hrs at 25 ℃.
Meanwhile, another sub-sample was kept at the same conditions without
fumigation. After CHCl3 was fully removed, organic C
from fumigated and unfumigated soil samples were extracted with 0.5Mol
K2SO4, with a soil: extractant ratio of
1:4 (w/v), and shaken at 150 rpm for 1 hr. Then, extractable organic C
in soil extracts was analyzed by a TOC analyzer (High TOC, Elementar)
after the filtration with Whatman no. 2v filter paper. The microbial
biomass C and N were measured as the difference between fumigated and
non-fumigated samples and normalized to the weight of the soil fraction.
(5) The PLFAs were extracted, fractionated, and quantified as described
by Bossio and Scow (1998). Frozen soil aggregate samples (equivalent to
8 g dry mass of soil) were extracted by a mixture of methanol,
chloroform (CHCl3), and phosphate buffer in a volumetric
ratio of (2:1:0.8) for 2 hrs. The sediment was extracted for another 30
mins. After the extractions, the supernatants were transferred to a
separation funnel and put to rest overnight. After the separation, the
CHCl3 layer was obtained and dried under
N2. Through the elution with CHCl3,
acetone, and methanol successively by silica-bonded phase columns, the
polar lipids were separated from neutral lipids and glucolipids (Supelco
Inc., Bellefonte, PA). By a mild alkali methanolysis, polar lipids were
converted into fatty acid methyl esters. The extractants were then
redissolved in 300 μL hexane which contained methyl nonadecanoate fatty
acid as an internal standard. Samples were analyzed by an Agilent 6850
gas chromatograph coupled with a flame ionization detector and a HP-5
capillary column (25.0 m × 200 μm × 0.33 μm). Peaks were identified with
a microbial identification system (Microbial ID. Inc., Newark, DE, USA).
In this study, fatty acids with percentages higher than 0.5% of the
total were considered. The i14:0, i15:0, a15:0, i16:0, i17:0, 14:1ɷ5c,
16:1ɷ7c, cy17:0, 17:1ɷ6c, 17:1ɷ8c, 18:1ɷ7c, cy19:0, 16:1 2OH and a17:0
were used as biomarkers for Germ, while the 18:1ɷ9c, 18:2ɷ6c, 18:3ɷ6,
16:1ɷ5, 10me16:0, 10me17:0 and 10me18:0PLFAs were used as biomarkers for
fungi (Zak et al. 1996; Pinkart et al. 2002; Zhang et al. 2015; DeForest
et al).
2.4 Statistical Analysis
All data were expressed as mean±1 standard error of mean(SEM). The data
for the 0–50-cm soil layer were used to analyze the C and N storage
potentials of the grassland. An analysis of variance (ANOVA) was used to
assess the effect of land-use change on C and N storage and
microbiological differences. All statistical analysis were performed
with the software program R(3.4.1). The meteorological data of this
study are sourced from China Meteorological Data Network
(http://www.worldclim.org). Sigmplot and R software were used for
mapping.
3. Results
3.1 Variate of Carbon and Nitrogen Pools
3.1.1 Soil Carbon and Nitrogen Pools
The values of the total soil C storage differed significantly among the
four sites (P < 0.01), varying from 0.4 g C
kg-1 for plot SD to 17.5 g C kg-1for plot NG. Similarly, the values of the total N storage differed
markedly among the four sites (P < 0.01), varying from
0.03 g N kg-1 for plot MD to 1.7 g N
kg-1 for plot NG (Fig. 1). The total C storage
decreased substantially with grassland degradation, and increased to a
significant extent with the introduction of NG. The C concentration in
soil was far higher in the 0–10-cm, 10–20-cm and 20–30-cm soil layers
than in other soil layers (Fig. 1a). The N concentration in soil was far
higher in the 0–10-cm and 10–20-cm soil layers than in other soil
layers (Fig. 1b).
Compared to SD, NG increased the total C and N storage in the 0–50-cm
soil layer by 97.3% and 98.1%, with an annual increase rate of 1.71%
and 1.72%, respectively. The total C and N storage increased
logarithmically with the duration of NG (P < 0.05)
(Fig.1).
3.1.2 Vegetation Carbon and Nitrogen Pools
The C and N stored in the above-ground biomass were less than 537.2 g C
kg-1 and 20.1 g N kg-1, respectively
(Fig. 2), the C and N stored in the roots were less than 503.5 g C
kg-1 and 16.5 g N kg-1,
respectively., and the C and N stored in the litter were less than 510.0
g C kg-1 and 16.6 g N kg-1,
respectively, accounting for negligible amounts (<1% of the
total) of total C and N storage in the ecosystem. The total C storage
(including C stored in above-ground biomass, litter, roots, and 0–50-cm
soil layers) differed significantly among the four sites (P<0.01), and the C storage varied remarkably among the
different pools (Fig. 1 and Fig.2). The amount of C stored in plants
accounted for over 90% of the total C storage, the C stored in soil was
very low (<10%), compared to other pools, and the amount of C
stored in the roots varied from 8.5 g C kg-1 for plot
SD to 432.1 g C kg-1 for plot NG. Similarly, the total
N storage (including N stored in above-ground biomass, litter, and
roots) differed significantly among different grasslands (P< 0.01). The total N storage varied from 8.5 g N
kg-1 for plot SD to 12.5 g N kg-1for plot NG (Fig. 2).
The C and N storage varied remarkably among the different pools at
different season, increase or decrease among the different pools (Fig.
2a and Fig.2b).
The N stored in litter and roots was very low, compared to that in
above-ground biomass (Fig. 2).
3.1.3 Microbial Biomass Carbon and Nitrogen Pools
The values of MBC storage differed significantly among the four sites
(P < 0.01), varying from 0.9 mg MBC
kg-1 for plot MD to 200.7 mg MBC
kg-1 for plot NG (Fig. 3a). Similarly, the values of
MBN storage differed significantly among the four sites
(P < 0.01), varying from 0.8 mg N kg-1for plot MD to 32.0 mg N kg-1 for plot NG (Fig. 3b).
The C stored in the MD were less than 26.9 mg MBC kg-1and 4.41 g MBN kg-1, respectively (Fig. 3a). The MBC
and MBN stored in the dry-season were less than 19.6 mg MBC
kg-1 and 6.25 mg MBN kg-1 among the
MD, respectively, while the MBC and MBN stored in the wet-season were
less than 26.7 mg MBC kg-1 and 2.2 mg MBN
kg-1among the MD, respectively. The total MBC storage
(including MBC stored in the 0–50-cm soil layers) differed significantly
among the four sites (P < 0.01). The total MBC storage
decreased substantially with grassland degradation, and increased to a
significant extent with the introduction of NG (Fig. 3).
The amount of MBC stored in NG accounted for over 70% of the total MBC
storage in the site, and the MBC stored in deep-soil layer was very low
(< 20%), compared to other layers. The amount of MBN stored
in the soil varied from 0.5 g MBN mg kg-1 for plot MD
to36.5 mg MBN kg-1 for plot NG.
The MBN concentration in soil was far higher in the 0–10-cm, 10–20-cm,
and 20–30-cm soil layer than in other soil layers (Fig. 3).
The total MBC and MBN storage varied remarkably among the different
seasons (P < 0.01), increase or decrease among the
different pools (Fig. 3a and Fig.3b).
Compared to NG, more than 90% of the soil microbial biomass in MD had
been lost. Compared to MD, NG increased the total MBC and MBN storage in
the 0–50-cm soil layer by 93.5% and 90.7%, respectively.
The total MBC and MBN storage did show a similar trend of increase with
the duration of NG (P <0.05) (Fig. 3).
3.1.4 Bacteria and Fungi Pools
The content of sand and gravel contributed more to soil fungi and had a
positive effect (Fig. 8). Microorganisms were not always limited by soil
C and N, and the bacteria and fungi storage did not show a similar trend
of increase with the duration of NG (P < 0.05). The
percentage of bacteria and fungi varied remarkably among the different
grazing pools (Fig. 4). The variation range of soil bacteria percentage
in dry-season was 26.5%~40.7% and wet-season
5.4%~47.0%, respectively, while the variation range of
soil fungi percentage in dry-season was 3.9%~10.42%
and wet-season 1.3%~8.9%, respectively.
In different desertification, the overall trend of soil fungi is that SD
has the lowest level, and MD and LD are similar (Fig. 4).
Our results partially supported a decreased contribution of fungi PLFAs
in wet-season compared to dry-season (in the 0-40-cm soil layers) and an
increased contribution of fungi PLFAs in wet-season compared to
dry-season (in the 40-50-cm soil layers) (Fig.4a and Fig.4b). With the
increase of precipitation, soil bacteria and fungi significantly
increased or decreased, which indicated that soil microorganism was
greatly influenced by precipitation patterns. Bacteria and fungi content
are found both highest in dry-season among MD and LD, while the highest
content of bacteria and fungi may be distributed in wet-season among NG.
Additionally, they are remarkably different from lowest to highest
content, SD< MD< LD< NG, in wet-season.
The distribution of soil bacteria in the surface layer was significantly
higher than that in the bottom layer (0-30cm) among SD, MD, and LD
(Dry-season, Fig.4a).
3.2 Relationship between Nutrient Pools and Grazing Intensity
3.2.1 Effects of Soil Aggregate Size on Grazing Intensity
Compared to SD, NG can increase silt and clay storage in the 0–50-cm
soil layer by 90.2% and 90.5% (annual increase rates), respectively.
In wet-season, changes in grazing management can lead to annual
increases of about 1.3% and 1.4% in silt (R2=0.97)
and clay(R2=0.88) storage, respectively. The soil silt
storage sustained an initial rapid increase with the introduction of NG,
followed by a steady phase of silt storage with grazing time (Fig. 6).
3.2.2 Relationship between Soil Nutrient Pools and Grazing Gradient
Compared to SD, NG can increase C and N storage in the 0–50-cm soil
layer by 97.3% and 98.1%, and the annual increase rates are both about
1.7% (R2>0.89;
R2≧0.98) for C and N, respectively. Moreover, the
total soil C content is, to a certain extent, dependent on the type of
land-use. The soil C and N storage sustained an initial rapid increase
with the introduction of NG (Fig. 6).
3.2.3Relationship between Plant
Nutrient Pools and Grazing
In our study, regression analysis indicated that there were less
relationship between plant nutrients (including C and N stored in
above-ground biomass, litter, and roots, R2< 0.5; in dry-season and wet-season, respectively) and grazing
exclusion (Fig. 5).
3.2.4 Relationship between Microbial Biomass Carbon and Nitrogen Pools
and Grazing
In our study, data showed that the soil MBC storage had decreased by
93.5%, 85.6%, and 84.7% for plot SD, MD, and LD
(R2=0.98, dry-season; R2=0.50,
wet-season), respectively, and the soil MBN storage had decreased by
90.7%, 95.2%, and 89.5% (R2=0.74, dry-season;
R2=0.50, wet-season), respectively (Fig. 6).
Therefore, a large amount of MBC and MBN has been lost in the last seven
decades across grasslands subjected to long-term heavy grazing. In
summary, the degradation of temperate grasslands due to long-term heavy
grazing has reversed the sequestration potential and led to MBC and MBN
loss by erosion and oxidation, instead of the C and N sequestration that
has been desirable in the past seven decades.
3.2.5 Relationship of Soil Fungi and Bacteria with Grazing Gradient
The change in land-use has no significant effects on bacteria and fungi
content in the grasslands of northern China. Particularly, low
correlation coefficients were observed in dry-season and wet-season
(R2<0.20;
R2<0.65) (Fig. 6). Enhanced diffusion rates
of nutrients with water addition may benefit bacteria more than fungi,
thereby decreasing the F/B ratio as nutrients are more efficiently
transported among the four sites.
3.3 Analysis of Driving Factors of Soil Degradation
Our PCA analysis showed that precipitation did not contribute much more
to the potentials of soil carbon and
nitrogen storage (including C and N stored in above-ground biomass,
litter, and roots). The soil silt and clay content may be considered to
make the modest contribution to the soil degradation. The second
contribution is considered to be the soil carbon and nitrogen storage.
The third is the microbial biomass, and the contribution of vegetative
sub-banks is greater than that of climatic conditions (Fig. 7 and Fig.
8).
Regression analysis showed that enclosure had a significant positive
effect on soil particulate matter composition (R>0.90;P <0.01). Soil micro-glasses contributed more than 70%
to soil nutrient pool (Fig. 6), and soil nutrient pool and microbial
pool increased functionally with grazing years (R>0.50;
P<0.05) (Fig.6, Fig.7, and Fig.8), indicating that soil has
strong dependence on and coexistence with disturbances, and soil quality
degradation is a synergistic reaction of grazing disturbance intensity.
Fencing is one of the most economical and effective measures for natural
vegetation restoration of desertified grasslands.
4. Discussion
4.1 Response of Soil Nutrient Pools to Grazing Gradient
The far-ranging employment of free-grazing as a land-use practice is
common in the temperate grasslands of northern China. A long period of
overgrazing has incurred the decline of grassland productivity,
deterioration of grassland, and the soil loss in vast areas (Dong and
Zhang, 2005). In the semi-arid grassland, the potentials of C and N
storage are approximately 17.5 g C kg-1 and 1.7 g N
kg-1, respectively. The productivity of grasslands
subjected to NG were stable or mature (Xiao et al., 1996; Bai et al.,
2004). Moreover, the result from a 17-year study (2000-2017) on MD
suggested that the semi-arid grasslands after more than 15 years of
grazing exclusion were very weak C source. It had been noticed that
after a long period of exclosure (i.e. >15 years), the C
and N storage was relatively higher. Seasonal dynamic of C and N storage
is not significantly different (P > 0.05). The same
phenomenon is observed in our study. One plausible albeit theoretical
account of it is that an increase in ANPP would drive greater
competition for all resources, some of which are nutrients and water,
and such an increased demand on nutrients would drive greater gross
rates of soil organic matter (Fig.1). In our observation, total C and N
were increased in wet-season (including above-ground biomass, ground
litter, and roots). This, similarly, would likely result in more soil
organic matter mineralization under natural disturbances, including
large-animal grazing. Through altering soil-water content, nutrient
availability and heterogeneity, productivity condition, and so forth,
grazing exerts its influences on the dynamics of C and N in a grassland
ecosystem (Hulbert, 1988; David et al., 1991; Collins and Smith, 2006;
MacNeil et al., 2008).
Despite these caveats, the estimation of the potential storage capacity
can help us systematically distinguish the effects of different
management strategies on the C and N storage of grasslands in northern
China. In this study, the value for site NG was about 537.2 g C
kg-1, in alignment with the previous estimate of
10–12 kg C m2 for the area (Wu et al., 2003), and it
was higher than the global mean value of 10.6 kg C m2(Post et al.,1982). Burke et al.(1995) have illustrated that a 50-year
period was enough for the recovery of active SOM and nutrient
availability. Thus, we suggest that a MD of at least a 20-year duration
would be reasonable for the restoration of the semi-arid grasslands from
a state of degradation in productivity, SOC and STN storage to a similar
one to natural grassland.
In this study, the soil C and N content in the surface layer is
relatively high, which is due to the fact that the surface soil can
adsorb newly imported organic matter quickly, thus inhibiting the primal
effect and soil organic carbon mineralization. The above findings are
similar to those in the early study (White et al., 1996).
4.2 Response of Soil Microbial Pools to Grazing Gradients
From the basic data, we observed that soil nutrient pools and soil
microbial biomass pools were similar in increase or decrease based on
our correlation analysis. The C and N storage varied remarkably with
significant positive correlation in the microbial biomass carbon and
nitrogen pools, respectively (Fig. 1, Fig. 3, and Fig. 4). And the PCA
analysis indicated that silt and clay were more important contributions
to the soil microbial biomass C and N storage than soil nutrient pools.
Contrary to the study of predecessors, we observed that higher water
availability did not play a positive role in soil nutrient pools and
micro organisms.
With regard to MBC and MBN storage, the grasslands with a higher
potential for MBC and MBN sequestration are those that have been
depleted by poor management strategies in the past. Based on our
findings, we conclude that the temperate grasslands of northern China
exhibit tremendous potential for the increase of their MBC and MBN
storage.
4.3 Effects of Precipitation on Carbon and Nitrogen Pool Cycling
This study showed minor effects of precipitation changes on the recovery
of soil nutrients and their microbial biomass content (Fig. 1, Fig. 3,
and Fig. 8). There is no significant effect with different precipitation
patterns (Ma et al., 2020), which may be led by a high recovery, often
with a low productivity at low N.
In grasslands with nutrient limits on productivity, the resistance of a
natural system to a disturbance decreased, whereas its
recovery/resilience increased with growing limited resource levels such
as N and P (De Angelis et al., 1989), which again indicated that given
limited nutrient resources, ecosystem resistance and recovery/resilience
could be inversely related (Herbert et al., 1999).
Our PCA analysis revealed that precipitation did not contribute much
more to soil carbon and nitrogen storage potential (including C and N
stored in above-ground biomass, litter, and roots) (Fig. 8). Therefore,
contrasting with resistance, the recovery of ANPP or plant growth after
a dry year or a drought event was often found higher when the
precipitation decreased, which was consistent with other previous
studies (Fig. 2 and Fig. 6) (e.g., Xu et al., 2009). Among various
ecosystems, the more severe and prolonged droughts are, the more time
the ecosystems need to recover (Schwalm et al., 2017). However, in the
SD, MD and LD experiment, a great increase in the recovery of nutrient
with dry-season regimes may be partly attributable to the fact that the
vegetation has not been permanently damaged, leaving the vegetation with
the potential to survive even under pre-drought conditions (Schwalm et
al., 2017). Additionally, it might be ascribed to the compensatory
growth and soil nutrients’ releases following rewetting (Mackie et al.,
2019). Compensatory growth may play a critical role in the rapid
recovery (Chen et al., 2020). As recently reported by Sankaran (2019),
in these arid and semi-arid savannas, the ability of the plant community
to recover from drought stress hinges on the length, severity and
frequency of the pre-drought period. Therefore, whether and how the
recovery is constrained by the pre-drought may depend on the severity
and duration of the pre-drought (Xu and Zhou, 2007) as well as the
growth stage of the plant (Sankaran, 2019). The results of current
studies show that there is no strong effect on soil resilience following
a change in rainfall regime (Fig. 1 and Fig. 2), which is consistent
with the result in a semi-arid grassland (Xu et al., 2014). This may be
due to the fact that the whole process of normal precipitation, less
precipitation, and rewetting cycle is covered by the resilience metric;
and an inverse correlation between resistance and recovery may indicate
that a trade-off occurs with resilience.
A previous study indicated that water addition significantly increased
fungal abundance in all soil aggregate classes (on average by 37.4%).
The composition of bacterial community is more affected by rainfall than
that of fungal community (Manzoni et al.2012). In our study, the changes
of bacteria and fungi variable by precipitation may be the effectiveness
of water in wet-season. Comparatively, the content of sand and gravel
contributed more to the growth of soil fungi and had a positive effect
(Fig. 6), and bacteria have been found to be better adapted to the
environment. Soil acidification is conducive to bacterial reproduction
but not to fungal (Pennanen et al. 1998). In this study, the same
conclusion can be drawn from the slow decreased change of bacteria with
altered rainfall patterns. Conversely, it has also been concluded that
fungi increased by as much as 80.8%, which could suggest that there
could be better adaption of fungi to lower soil pH than that of bacteria
(Wang et al. 2004 ). With the increase of rainfall in the growing
season, the growth rates of bacteria and fungi are different, indicating
that water has different effects on different microbial groups,
consistent with the results of Zhang et al. In addition, higher water
availability could further improve substrate diffusion and nutrient
accessibility to soil microorganisms, thereby advancing microbial growth
and increasing total PLFA concentration and abundance of individual
microbial communities (Dungait et al. 2012; Nielsen and Ball 2015).
We observed that given added water, there was a significant decrease in
fungal abundance (Fig. 3b) and F/B ratio, indicating a change in the
microbial diversity with enhanced bacterial proliferation over fungi
under favorable water conditions (Griffiths et al. 1998). This may be
ascribed to the different growth strategies of fungi and bacteria
(hyphal growth versus individual cells) (Frey et al. 2004). Fungal
hyphae can transport nutrients and resources from one microsite to other
sites where nutrients are limiting their growth (Strickland and Rousk
2010). However, increased diffusion rates of nutrients with added water
may be more beneficial to bacteria than to fungi, thus decreasing the
F/B ratio as nutrients are translocated more efficiently within the
three soil fractions (Dungait et al. 2012).
4.4 Response of Plant Nutrient Pools to Grazing Gradient
Recent grassland-related field studies in Inner Mongolia showed that a
plant NPP of about 1.5 tons ha-1 was limited by both N
and water as addition above 5.25-17.5 g N m2yr-1 of background increased NPP by 13%-62% (Bai et
al., 2010), whereas water addition increased above- and below-ground NPP
by 32.9% and 38.3%, respectively (Xu et al., 2010). However, soil
microorganisms are not limited by the same factors that constrain plant
systems (Hobbie et al.,2005; Wei et al., 2013). For instance, the study
by Wei et al. (2013) showed that there were differences in N saturation
levels (threshold levels for N demand) between plants and soil
microorganism, emphasizing that microbes could be limited by C or P
while plants were limited by N (Treseder, 2008). In addition, under
higher N availability in temperate grasslands, both the size and
activity of soil microbial biomass were found decreased (Gutknecht et
al., 2012; Wei et al., 2013).
Plant nutrients are not limited by soil nutrients and microbial
conditions, which may be due to the time and space lag of plant
succession compared with soil nutrients. Vegetation replacement and
nutrient status change require a long buffer period.
Soil organic carbon content decreased gradually from the NG to the SD,
but the plant performance was not synchronous. Soil’s response to
desertification is more sensitive than vegetation, and the change of
plant nutrients has a certain lag in time.
4. 5 Effects of Enclosure on Carbon and Nitrogen Pools
Regression analysis showed that enclosure treatment had a significant
positive effect on soil aggregates composition and nitrogen pools
(R>0.90; P<0.01). The soil macro-aggregates
contributed more than 70% to soil nutrient storage (Fig. 6). Soil
nutrient and soil microbial pools increased functionally with grazing
year(R>0.50;P<0.05)(Fig. 6). The recovery of
the heavy-grazing area to a stable and healthy natural grassland
requires more than 50 years of enclosure management (Fig.6). A large
amount of evidence from previous studies suggested that a period of 10
years after enclosure of desertified grassland was a process of soil
development from quantitative to qualitative change, and organic matter
content reached 2.86% after 17 years of containment, 36.7 times of the
initial containment. This study showed that there was a high correlation
between soil nutrients and microbial nutrients pools with different
grazing levels in the fenced time, and the longer control time is, the
more annual nutrients return to the soil storage. And the number of soil
surface bacteria in the initial stage of fenced grassland is
43.348*104 soils/g. After 17 years of containment,
975.51 *104 soils/g were observed. The increase is
40.95 times of the initial containment. For SD and MD grassland, the
restoration process may be promoted by reseeding of superior native
plant seeds under proper organic fertilizer input and enclosure.
5. Conclusion
Land-use change has significant effects on C and N storage in the
grasslands of northern China. The storage of C and N has decreased
greatly because of grassland degradation led by long-term heavy grazing.
The storage potentials of C and N in the semi-arid grasslands are
approximately 537.2 g C kg-1 and 16.6 g C
kg-1, respectively, so, there is huge potential for
the increase of C storage in the temperate grasslands of northern China
by improvement of grassland use or management. Micro-aggregates
availability was suggested to be the main limiting factor of both NPP
and microbial biomass C and N storage in this semi-arid grassland soil.
Moreover, it is found that the site with 17 years of fencing has the
highest level of soil micro-aggregates. Fencing is the most economical
and effective measure for natural restoration of degraded grassland and
the restoration of heavily grazed areas to stable and healthy takes at
least 50 years.