Changes in Soil organic carbon and total nitrogen following long-term grazing exclusion in Saral grasslands, Kurdistan (Iran)

doi:10.21203/rs.3.rs-2500894/v1

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Changes in Soil organic carbon and total nitrogen following long-term grazing exclusion in Saral grasslands, Kurdistan (Iran)

https://doi.org/10.21203/rs.3.rs-2500894/v1

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Background and aims: Land use change favouring agriculture is considered a global threat as it interrupts ecosystem services of grasslands. However, exclosure from grazing has been widely known as effective in maintaining or even restoring grasslands. This study examined the impacts of long-term (45 years) exclosure, heavy grazing and rainfed wheat farming on soil organic carbon (SOC) and total nitrogen (TN) in the cold semi-arid grasslands of Saral Agricultural and Natural Resources Research Station, Kurdistan, Iran.

Methods: In this investigation, 90 soil samples were taken from 0-20 cm depth within an area of 0.5 hectare for each land use (exclosure, heavy grazing, wheat farming). Samples were collected in the four non-consecutive years of 2009, 2013, 2017 and 2021. The Walkley-Black and Kjeldahl methods were used to measure SOC and TN respectively.

Results: Results showed that value of SOC differed significantly among different land uses (P<0.01). In the studied years, SOC in the excluded site in 2021 had the highest level with 34,787 kg/ha, while the grazed site recorded the lowest SOC with 26,342 kg/ha in the same year. TN was also the highest at 0.130 percent in the excluded area in 2021, while TN was lower at 0.102 and 0.103 percent in the grazed and wheat-farming areas respectively in 2021.

Conclusion: This suggests that in grasslands with the same ecological conditions, managers can choose appropriate grazing intensity as an alternative to either heavy grazing or rainfed wheat farming, though more investigations on the efficiency of different grazing intensities are needed.

heavy grazing

exclosure

grazing intensity. soil organic carbon

soil nitrogen

rainfed wheat farming

Grassland is the largest terrestrial biome, covering 3.2 billion ha worldwide, and a large part of this area is used by grazing herbivores (Scholtz and Twidwell 2022). Semi-natural grasslands are traditionally managed under diverse management regimes and host a great number of plant species, which is essentially why they contribute to the biodiversity of agricultural landscapes (Walden 2019). As one of the most widespread terrestrial ecosystems in the world, grasslands play a crucial role in the global carbon (C) cycle and provide key ecosystem services. Given that atmospheric carbon dioxide has dramatically increased over the last century, it has become increasingly crucial to know and understand the global carbon cycle (Schlesinger and Bernhardt et al. 2013). Soils are the largest terrestrial pool of carbon (C), encompassing about 1500 Pg organic C in the top meter worldwide. Therefore, small changes in the soil organic carbon (SOC) reservoir may significantly influence the concentration of carbon dioxide (CO2) in the atmosphere. With rising levels of atmospheric CO2, there has been growing interest in SOC sequestration in recent years. It is estimated that grasslands store 10–30% of the global SOC with a C sequestration rate of 0.5 Pg C year − 1 (Batjes 2018). Due to the strong potential for C sequestration, grassland soils may act as a sink for atmospheric CO2 and contribute to the mitigation of climate change.

Management strategies play a pivotal role in nutrient concentration in soil and plants: many ecosystem services thus depend largely on the management and intensity of grassland use (Hein 2006; Eldridge et al. 2016; Lu et al. 2017; Zhao et al. 2021). Grazing by livestock is the most common management strategy which is applied in managing grasslands globally. Several researchers have shown that intensity of grazing affects not only forage quality but plant species composition (Skinner et al. 2004). Grazing also alters physiochemical properties of soils, in particular carbon and nitrogen cycles (Ren et al. 2012). A moderate grazing intensity can result in a heterogeneous sward structure with spatial variations in height, which in turn affects the floristic composition and structure of plant communities (Karami et al. 2019). Hence, management strategies applied in grasslands affect plant traits, and other studies have demonstrated that grazing intensity regulates the biogeochemical cycles of C, N and P in grassland ecosystems through affecting plant nutrient use efficiency and soil physicochemical processes (He et al. 2019). For example, Zhu et al. (2021) showed that grazing prohibition not only increased the aboveground biomass, total biomass and total meadow coverage but increased the soil total carbon, soil organic carbon, soil total nitrogen, soil total phosphorus and soil available phosphorus in subalpine meadow and swamp meadows. Moreover, grazing, as a key anthropogenic disturbance, affects ecosystem C, N and P cycles as well as their stoichiometry (He et al. 2020). In addition to SOC, soil N in grasslands has received much attention because it plays a key role in global biogeochemical cycles. Also, the interactions of C and N in soils are of great importance for regulating the main ecological processes of nutrient cycling and energy flow.

Land use change of forests, rangelands and grasslands to agricultural lands is the most widespread conversion globally with dramatic changes in soil quality, soil productivity, vegetation type, carbon sequestration, microbial community composition and decreases in microbial biomass. Those alterations have crucial impacts on the global nitrogen cycle with subsequent effects on climate change and degradation of biodiversity (Tecimen 2017). Since most of the uncultivated grasslands are grazed by large mammals, the effect of grazing on SOC is vital for C sequestration in grasslands. Unfortunately, large areas of grasslands are experiencing overgrazing worldwide, which not only reduces plant biodiversity and productivity but also modifies the structure and function of the ecosystem, leading to a depletion of SOC stock (McCherry and Ritchie 2013; Liu et al. 2022). Changes in land use also alter the C and N cycles, which may change C and N accumulation in the soil (Piñeiro et al. 2006). In the context of climate change, the restoration of overgrazed and degraded grasslands is thus urgently needed to recover SOC levels. Research by Abdalla et al. (2022) has demonstrated that grassland restoration as a soil management strategy increases SOC in grasslands.

Saral’s semi-natural grasslands in Kurdistan provide habitat for wild animals and endemic plant species and for recreational opportunities for people in the region (Karami et al. 2021). These grasslands also feed numerous springs and water resources which are used for agricultural, recreational, and household purposes. Although previous studies reported on the effects of grazing intensity on plant traits, there is no clear answer to the question of whether there are significant differences between rainfed wheat farming, heavy grazing, and long-term grazing exclosure with regard to SOC and TN concentrations in semi-natural grasslands. More importantly, there is insufficient evidence about changes in amounts of SOC and TN over time. This investigation, thus, was undertaken in 2022 using samples collected in 2009, 2013, 2017 and 2021 in the Saral grasslands (Kurdistan, Iran) with the objective of comparing the effects of the different land uses strategies of heavy grazing, exclosure from grazing and rainfed wheat farming on the amounts of SOC and TN.

Study area

This investigation was carried out in the 4 non-consecutive years of 2009, 2013, 2017, and 2021 in the Saral Agricultural and Natural Resources Research Station approximately 45 km from Sanandaj (46°46' to 47°47' E and 35°39'N to 35°40'N latitude) at an altitude of 2090 m (Fig. 1). The excluded, grazing and rainfed wheat farming sites have the same general conditions with regards to topography, climate, and soils. Mean annual precipitation is 473 mm. In most years the maximum monthly precipitation occurs in February and the minimum in June. The mean annual temperature is 16°C and the study area is described as cold semi-arid (BSk) in the Köppen Climate Classification. Soils have a sandy loam texture, with pH 6.5 in the grazed and rainfed wheat farming areas and pH 7 in the excluded sites. The private (governmental) excluded area which has a research use was excluded from grazing by fencing constructed around the site in 1976. Grazing (grazed sites) by sheep, goats, cows, and wild animals takes place from late April to late July without any additional management practices; overall grazing intensity has been determined as heavy (Karami et al., 2019). At the grazed sites, wild and domestic animals have had continuous access since 1976.
In Western Iran, the main wheat variety grown in the rainfed farming areas is Sardary (Tavakoli et al. 2013).

Species in grazed, excluded and rainfed wheat farming sites

Overall, 58 plant species were found, of which 52 species were recorded in the excluded sites and 47 in the grazed sites (Table 1). Most species thus occurred in both the grazed and excluded areas, but the excluded sites recorded 6 more species.

Table 1 Names and presence (+ is present, - is absent) of the species in grazed, excluded and rainfed farming sites.

Species	Grazed Site	Excluded Site	Rainfed farming
Achillea millefolium L.	-	+	Wheat
Achillea vermicularis Trin.	+	+
Alyssum linifolium Steph.ex Willd.	+	+
Astragalus apricus Bunge	+	-
Astragalus geminanus Boiss. & Hausskn	+	-
Astragalus tortuosus DC.	+	+
Bellevalia glauca (Lindl.) Kunth	+	+
Bromus danthoniae Trin.	+	+
Bromus tectorum L.	+	+
Bromus tomentellus Boiss.	-	+
Bunium cylindricum (Boiss. & Hohen.) Drude	-	+
Campanula involucrata Auch.ex DC.	+	+
Centaurea behen L.	+	+
Centaurea pseudoscabiosa Boiss & Buhse.	+	+
Cephalaria syriaca (L.) Roemer & Schultes	+	+
Cerastium inflatum Link ex Desf.	+	+
Chaerophyllum macropodun Boiss.	+	+
Chenopodium album L.	+	+
Cruciata taurica (Pallas ex Willd.) Ehrend.	+	+
Dactylis glomerata L.	+	+
Echinops haussknechtii Boiss.	+	+
Elymus hispidus (Opiz) Melderis	+	+
Eremopoa persica (Trin.) Roshev.	+	+
Eryngium thyrsoideum Boiss.	+	+
Euphorbia aleppica L.	+	+
Ferula Haussknechtii Wolff ex Rech. f.	-	+
Festuca ovina Boiss	-	+
Geranium tuberosum L.	-	+
Gypsophila bicolor (Freyn& Sint.) Grossh.	+	+
Heteranthelium piliferum (Banks & Soland.) Hochst.	+	+
Hordeum bulbosum L.	-	+
Hypericum scabrum L.	+	+
Isatis kotschyana Boiss. & Hohen.	-	+
lactuca serriola L.	-	+
Lallemantia iberica (Stev.) Fisch. & C.A. Mey.	+	-
Lathyrus chloranthus Boiss.	-	+
Myosotis lithospermifolia (Willd.) Hornem.	+	-
Neoa mucronata (Forsk.) Aschers.	+	+
Nonnea hypoleia Bornm	+	+
Onobrychis major (Boiss.) Hand-Mzt.	+	-
Onopordon acanthium L.	+	+
Onosma microcarpum DC.	+	+
Phlomis olivieri Benth.	+	+
Poa annua L.	-	+
Poa bulbosa L.	+	+
Ranunculus aucheri Boiss.	+	+
Rheum ribes L.	+	+
Rumex scutatus L.	+	+
Salvia multicaulis Vahl	+	+
Scariola orientalis (Boiss.) Sojak	+	-
Scilla persica Hausskn.	+	+
Scorzonera calyculata Boiss.	+	+
Silene commelinifolia Boiss.	+	+
Stroganowia persica Busch	+	+
Thymus kotschyanus Boiss. & Hohen.	+	+
Tragopogon bornmuelleri Rech.f.	+	+
Veronica anagalis- aquatica L.	+	+
Ziziphora clinopodioides Lam.	+	+

Field design and statistical analysis

In 2009, three land uses were selected which were located in close proximity to each other. To remove distortion from edge effects, a buffer zone of 30 m was retained between land uses. An area of approximately 0.5 ha was selected for each land use, with selected areas being identical in terms of environmental characteristics. Three transects (or plots) of 5 x 40 meters were laid out in each of the three land uses, giving a total of 9 transects. In each transect, 10 (1 × 1 m) plots were randomly selected as sub-plots. Soil samples were taken from a depth of 0–20 cm in the sub-plots using an auger (with a diameter of 6.1 cm) Sampling was done in the first week of each of the four years. Then, to obtain a composite sample that represented each transect, the soil from 10 sub-plots was thoroughly mixed. Each land use was thus represented by three composite samples. The collected soil samples were first air-dried and plant residues including plant roots and other visible residues were removed from the soil samples. Samples then were passed through a 2 mm sieve.

To measure soil organic carbon and total nitrogen, the Walkley-Black Method (1934) and Kjeldahl methods were used respectively. The hydrometric method (Bouyoucos, 1962) was used to measure soil particle size percentage (sand, silt, and clay).

Bulk density of the soil was measured by the cylinder method (Blake and Hartge, 1986) and EC meter was used to measure EC (Electrical conductivity) (Carter and Gregorich, 2007). Data for SOC and TN were also collected in the 3 non-consecutive years of 2013, 2017 and 2021 in the same way as in 2009, though soil particle size, bulk density and EC were measured only in the first year.

This research was conducted as a 2-level factorial experiment in a completely randomized design with three replications. The first factor was land use in three levels (heavy grazing, excluded and rainfed wheat farming) and the second factor included year in four levels (2009, 2013, 2017 and 2021). One way analysis was used to analyze the data of pH, electrical conductivity, percentage of sand, silt and clay, and bulk density. Analysis of Variance was done using SPSS software. To compare the data, Duncan's test was used at the five percent probability level.

Results of analysis of variance of pH, electrical conductivity, percentage of sand, silt and clay, and bulk density in the excluded site, grazed area and rainfed wheat cultivation in the study area are shown in Table 2. The results showed that the values of bulk density (P < 0.05), electrical conductivity, and percentage of sand and silt (P < 0.01) were significantly different among the three studied land uses, while there was no significant difference in clay percentage (P > 0.05).

Table 2

Results of analysis variance of pH, electrical conductivity, percentage of sand, clay, silt and bulk density in the investigated land uses.
Variable	Df	Mean Square	F	Sig.
pH	2	0.168^ns	1.716	0.271
EC (ds/m)	2	0.009^**	40.95	0.001
Sand (%)	2	102.750^**	16.54	0.006
Silt (%)	2	70.083^**	30.91	0.002
Clay (%)	2	10.433^ns	1.918	0.241
Bulk Density(g/cm³)	2	0.129^*	7.728	0.030

*and **: significant at P< 0.05 and < 0.01, respectively; ns = non-significant

Table 3 shows the average factors of pH, electrical conductivity, percentage of sand, clay, silt, and bulk density as well as soil texture in the three studied land uses. No significant difference between uses was recorded for pH, but electrical conductivity in the excluded rangeland was significantly higher than for the other two uses. Among the parameters that determine the soil texture, i.e., the percentage of sand, silt and clay, the amount of sand in the land under wheat cultivation is significantly higher than for the other two land uses, and the amount of silt is significantly less. There was no significant difference in the percentage of clay between the three studied land uses. Although the three land uses were significantly different in terms of the percentage of sand and silt, in general, the soil texture was the same (Table 3). Another physical factor which was studied was the bulk density. The results showed that there is a significant difference in bulk density between the three uses, with wheat cultivation areas (1.22 g/cm³) being lower compared to the other two uses of exclosure from grazing (1.59 g/cm³) and livestock grazing (1.67 g/cm³) (Table 3). There was no significant difference between the two land uses of exclosure and livestock grazing in terms of bulk density.

Table 3

Comparison of the average factors of pH, electrical conductivity, percentage of sand, clay, silt and bulk density as well as soil texture in the three land uses studied using Duncan's test.
Variable	Land Use
Variable	Exclosure	Grazed	Wheat-land
pH	6.9 ± 0.33^a	6.8 ± 0.01^a	7.3 ± 0.49^a
EC (ds/m)	0.412 ± 0.021^a	0.312 ± 0.006^b	0.326 ± 0.001^b
Sand (%)	19 ± 2.08^b	22.4 ± 0.87^b	31.9 ± 0.5^a
Silt (%)	25.66 ± 0.88^a	25.66 ± 0.88^a	16 ± 1^b
Clay (%)	55.33 ± 1.2^a	51.93 ± 1.7^a	52.1 ± 0.5^a
Bulk Density(g/cm³)	1.59 ± 0.04^a	1.67 ± 0.10^a	1.22 ± 0.5^b
Soil Texture	clay	clay	clay

Superscripts with the same letter show that there is no significant difference

Superscripts with a different letter show a significant difference

Effects of land uses and year on SOC concentrations

The results of the analysis of variance of SOC shows that there was a significant difference between the types of land use (P < 0.01) (Table 4). However, there was no significant difference between the year and the interaction between year × land uses (P > 0.05). The comparison of mean results for SOC using the Duncan test showed that SOC of the three different land uses for each studied year was not significantly different (Table 3). In 2009 mean results for excluded grasslands were significantly different, while no significant difference was seen in figures for the grazed site and wheat farming. SOC in the exclosure site (33,152 kg/ha) was substantially higher than the 26,947 kg/ha recorded in 2009 for the grazed area, and the 24,864 kg/ha for rainfed wheat farming. In 2013, although the differences were not as great as in 2009, the excluded grasslands again had the highest proportion of soil organic carbon (Table 5).

The results of a comparison of mean soil organic carbon in 2017 and 2021 showed that the excluded grasslands had significantly higher levels of SOC than the land uses of rainfed wheat and grazed grasslands. In 2017, SOC in excluded area with 34563 was the highest whereas this figure in grazed area remained unchanged as 2013 at 27843kg/h. As for wheat farming, although SOC ratio was higher than 2013, the increase was not significant.

In 2021, SOC in excluded area reached to 34787 kg/ha and formed the highest SOC ratio among the studied land uses and the years of the study. While SOC ratio weight in the grazed area decreased over the years of the study, SOC in wheat farming increased over the time, though this increase was not significant over the last decade of the study.

Table 4

Analysis variance of soil organic carbon in three different land uses over 45 years
Source	Df	Mean Square	F	Sig.
Year	3	0.004	0.117	0.949
land use	2	0.460	14.96	0.0001
year * land use	6	0.003	0.113	0.994
Error	24	0.031
Total	36

Table 5

Soil organic carbon (kg/ha) in different land uses in studied years
year land use	2009	2013	2017	2021
Exclosure	33152^ab	33600^ab	34563^a	34787^a
Grazed	26947^bc	27843^abc	27843^bc	26342^bc
Wheat-land	24864^c	25984^bc	26790^bc	26432^bc

Superscripts with the same letter/s show that there is no significant difference

Superscripts with different letter/s show a significant difference

Effects of land uses and year on TN concentrations

Results of the analysis of variance of soil nitrogen shows a significant difference between the type of land use and associated soil nitrogen (P < 0.01) (Table 6). However, there was no significant difference between the year and the interaction between year and land uses (P > 0.05). Figure 2 shows that the difference between the three land uses was not statistically significant in terms of soil nitrogen in the years of 2009 and 2013. However, in 2017 and 2021, soil nitrogen concentrations were higher in the excluded grasslands than in the other two land uses of grazing and rainfed wheat growing, which were not significantly different from each other.

Table 6

Analysis of variance of soil nitrogen in three different land uses over 45 years
Source	Df	Mean Square	F	Sig.
year	3	0.000092	0.669	0.579
land use	2	0.002	12.998	0.001
year * land use	6	0.000046	0.338	0.910
Error	24	0.000
Total	36

Table 7

C/N ratios in different land uses
year land use	2009	2013	2017	2021
Exclosure	12.63	11.78	12.03	12.10
Grazed	11.27	11.73	11.34	11.34
Wheat-land	11.48	10.57	11.57	11.57

The present study provides a baseline to monitor the responses of SOC and TN to different land uses as recorded over a 45-year period in the cold semi-arid grasslands of Saral. The investigation was conducted in 4 non-consecutive years in specific local areas and is thus indicative of SOC and TN conditions of the Saral grasslands under three different land uses. Based on the results, heavy grazing changed plant species composition by replacing highly palatable grass species with undesirable plants (Table 1). These findings are in line with those reported by Karami et al. (2019) in which continuous grazing caused some species, namely Ferula haussknechtii and Prangos ferulacea, to disappear in a heavily grazed area. In our study (Table 1), fewer species were recorded in the grazed site compared with the excluded site. Fewer species in the overgrazed site has led to lower SOC relative to the excluded site. As reported in Table 5, exclosure from grazing significantly increased the SOC, while both overgrazing and rainfed wheat farming decreased SOC. These results support the findings of Su et al. (2003), Wachiye et al. (2022) and Abdalla et al. (2022) in which exclosure from grazing significantly increased plant-soil system C storage, and thus sequestration of atmospheric CO2.

The increase in SOC after grazing exclusion is attributed to higher vegetation cover and biomass, and higher soil water content (Wachiye et al., 2022). Abdalla et al. (2018) also showed that low grazing intensity significantly increased SOC for C4-dominated grassland which is associated with significant increases in TN and reduced soil bulk density. However, in humid subtropical grasslands, Wade et al. (2021) provided evidence that grazing increased soil carbon stocks in the top 0–5 cm layer, but no significant difference in SOC was recorded between grazed and ungrazed areas for the upper 0–30 cm soil depth.

The decrease in SOC in heavily grazed areas can be attributed to the decreased amount of plant material available to the soil due to plant removal by livestock grazing (Zhau et al. 2017; Liu et al. 2022). The grazing regime alters ecosystem carbon and nitrogen cycles through livestock feeding, trampling, and manure return in grassland ecosystems (Mipam et al. 2021; Yang et al. 2022). Plant defoliation and removal of different plant parts by livestock grazing are likely to decrease root elongation and biomass due to the decreased C allocation to roots (McSherry and Ritchie 2013). Grazing at higher intensities decreases aboveground biomass and plant production and this leads to a reduction in litterfall and litter mass (Liu et al. 2020; Wachiye et al. 2022). As Wachiye et al. (2022) and Li et al. (2018) have shown, a lower degree of vegetation cover in overgrazed areas could reduce root biomass thus lowering root respiration (an important source of carbon and energy for soil microorganisms). Also, changes in litter and root biomass decreased the soil bacterial community and diversity and then the microbial biomass and soil C pool (Bai et al. 2015a). It has been demonstrated that C inputs from root production and biomass may have a significant effect on soil C storage compared to aboveground biomass (Lu et al. 2017).

Along with grazing intensity, the type and size of the grazers browsing a site is likely to be important in affecting carbon and nitrogen cycles, though the probable effects are more indirect. For example, lower SOC in lightly grazed sites has been attributed to major defoliation and massive trampling by big mammals including buffaloes and elephants (Amra et al. 2020; Wachiye et al. 2022). This shift in plant species can affect the soil organic matter and root biomass (Klumpp et al., 2009).

Consistent with the results of An and Li (2014), the highest TN concentrations in our study were recorded in excluded areas compared to the rainfed wheat farming and overgrazed sites (Fig. 2) – that is, compared to the other two land uses, grazing exclosure significantly increased TN. He et al. (2018) also demonstrated that light grazing significantly increased the C:N, C:P and N:P ratios of soil but heavy grazing decreased these ratios. Lower TN concentrations in overgrazed areas occur from frequent livestock trampling. Soil compaction and less plant productivity then result in higher bulk density and crust formation, with crust formation subsequently leading to soil vulnerability to water erosion and C and N losses (He et al. 2022; Ahmad et al. 2022). However, at higher grazing intensities, inputs of livestock urine and faeces can partly compensate for N loss which will lead to a decreased C:N ratio (He et al. 2022). There was a slight difference in TN amounts between the rainfed wheat farming and the overgrazed site (Fig. 2), which can be explained by grazing-induced root exudation, along with input of livestock urine and faeces. In rainfed wheat farming, however, lower TN is attributed to removal of residue, stubble burning and plowing twice a year (Chan and Heenan 2006). Using fertilizers in rainfed wheat farming compensates for the lack of soil fertility. However, in the studied area economic conditions of local farmers over the last few decades have made the cost of fertilizers prohibitive, so that TN follows a decreasing trend from the beginning of the period. Although the amount of SOC and TN changed within different studied land uses, the difference between the year, and the interaction between year × land uses, was not significant- that is, SOC and TN sequestration in different land uses remained substantially the same over the study period. With regard to the effects of time on changes in the SOC and TN values, our results do not support the findings of Wang et al. (2016) in which SOC and TN decreased over time. They demonstrated that SOC and TN values were higher in the first year of their 5-year experiment.

Our results contribute to closing the existing knowledge gap regarding the effects of different land uses undertaken in semi-natural grasslands. This information is of great importance in understanding carbon cycling in semi-natural grassland, as well as the identification of the potential consequences of increasing land pressure caused by rising livestock numbers. Of the three land uses investigated here, areas excluded from grazing had consistently higher SOC and TN than heavily grazed or rainfed wheat-growing areas. In addition, these levels of SOC and TN did not decline over the four non-consecutive years of study in non-grazed lands and can therefore act as a ‘benchmark’ for other land uses. More research is needed (a) on evaluating the optimum grazing intensity for soil C and N storage in semi-arid regions, and (b) on developing methods for enhancing C storage and maintaining N levels within the constraints of existing rainfed wheat-farming systems.

Acknowledgement: We hereby thank the staff at the University of Kurdistan for their time and cooperation.

Funding: This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors.

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Changes in Soil organic carbon and total nitrogen following long-term grazing exclusion in Saral grasslands, Kurdistan (Iran)

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