Change in Land Use Affects Soil Organic Carbon Dynamics and Distribution in Tropical Systems

Abstract

Anthropogenic land cover change is directly responsible for the deforestation and degradation of tropical forests. In this context, assessing soil organic carbon (SOC) stocks is key to understanding the impact of anthropogenic activities on SOC so that we can implement management practices that effectively reduce emissions or promote carbon sequestration. Our objective was to assess the effect of land-use change on the dynamics and distribution of SOC in three systems (agriculture, pasture and agroforestry) after 40 years of deforestation in a tropical dry forest in the central–eastern region of Honduras. For this purpose, the bulk density, percentage of coarse fragments (>2 mm) and soil organic carbon content were determined at three depths (0.00–0.10 m, 0.10–0.20 m and 0.20–0.30 m). The results showed an increase in bulk density for all new uses, although soil compaction had not yet occurred. In terms of total soil organic carbon (TOC) stocks, deforestation caused a decrease from 17% to 48% in agricultural and agroforestry soils, respectively; on the other hand, grasslands did not show significant differences compared to tropical dry forest, suggesting that they have a high potential as carbon sinks in deforested tropical areas. However, this did not imply a better state of the system, as the greatest increases in bulk density were found in pastures.

1. Introduction

Ensuring food security and mitigating the effects of anthropogenic global warming are among the most pressing challenges facing humanity today. These changes have led to climatic changes that threaten both food production and the well-being of future generations. As a result, we are facing a triple crisis: climate change, pollution and biodiversity loss [1,2].

According to the United Nations [1], approximately 94% of global deforestation is due to agricultural (50%), livestock (38%) and urban (6%) expansion. Agriculture is responsible for more than a third of greenhouse gas (GHG) emissions, 70% of freshwater use and terrestrial biodiversity loss, and it is directly responsible for soil degradation, with an estimated 20–40% of Earth’s land surface showing varying degrees of degradation [3]. Livestock accounts for 12% of anthropogenic greenhouse gas emissions and 40% of total emissions from agrifood systems [4].

The United Nations Organization, in the 2030 Agenda for Sustainable Development Goals (SDGs), promotes the protection, restoration and sustainable use of terrestrial ecosystems and the sustainable management of forests, as well as combating desertification, reversing land degradation and halting biodiversity loss [1]. Therefore, various nature-based strategies (NBS), referred to as natural climate solutions, have been proposed in conjunction with forest restoration, rangeland restoration and the adoption of climate-smart agricultural practices [5].

The Intergovernmental Panel on Climate Change (IPCC) assigns an important role to soil organic carbon (SOC), considering increasing soil carbon sequestration capacity and soil conservation to avoid losses as strategies to combat climate change [6]. On a global scale, soils are the second largest C sink in nature, surpassed only by the oceans. They store about 1500 Pg at a depth of 1 m, followed by the atmosphere (880 Pg) and vegetation (620 Pg) [7]. The uptake and storage of C in soils depends on defining, limiting and reducing factors. Defining factors are related to the parent material, geomorphologic terrain characteristics (slope and aspect) and soil mineral composition, which determine the texture, depth, bulk density, rock fragment content, drainage and erodibility [8]. Limiting factors include net primary production, forest composition and climate (temperature and humidity). Reducing factors are related to erosion, deforestation and land-use processes [9].

In turn, deforestation and changes in land use and management can alter the amount of carbon stored in the soil [10]. In developing countries, forest degradation and deforestation, together with the expansion of agriculture, transport, industry and the domestic sector, are accelerating land-use change and altering the carbon balance. In addition, globalization processes and population growth have driven the expansion of the agricultural frontier, increasing land use for agricultural activities. This has led to a deterioration in the physical, chemical and biological properties of soil, resulting in progressive soil degradation [11,12].

According to Díaz-Zorita and collaborators, the use and management of soil can significantly influence its characteristics, which would be the anthropogenic forming factor, since intensive activities degrade the soil by reducing or eliminating its cover, decreasing the stability of aggregates and the amount of SOC, thus favoring the development of soil degradation processes and, consequently, decreasing its quality [13,14].

In tropical and subtropical regions, SOC decline is closely linked to soil and environmental degradation. In recent decades (1990–2020), about 420 million hectares of forest have been lost worldwide, with tropical forests being the most affected [15]. These forests are important reservoirs of biodiversity and livelihoods and store about 68% of the global forest carbon. Of this, 45–55% is stored as vegetation biomass, and the remaining 10–30% is stored in the soil [6]. It is therefore foreseeable that increased demand for land for crops and pastures in tropical regions will put further pressure on these forests [16].

Regarding the dynamics and distribution of SOC content, Hairiah et al. [17] state that these contents generally tend to decrease with depth and that the most significant changes in reserves occur in the first 0.20 m to 0.30 m, with estimates of losses of 50 to 75% in highly degraded soils and 25 to 50% in less degraded soils [8]. Pellat et al. [18] also indicate that soils occupied by agroecosystems (grasslands, shrublands, agriculture and other uses) contain about 40% of all terrestrial carbon.

In Central America, the increase in GHG emissions due to land-use change is motivated by deforestation for agro-livestock, industrial and urban expansion. In this context, we hypothesized that land-use change and the introduction of managed grazing systems would maintain TOC stocks, while agroforestry systems and cultivated areas would decrease TOC stocks. Therefore, our objectives were the following: (i) to analyze the effects of deforestation and agricultural expansion on TOC stocks under pastures, agroforestry systems, and agricultural areas in comparison with an undisturbed tropical dry forest; (ii) to evaluate how land use affects TOC at a 0.30 m depth; and (iii) to determine whether the soil acts as a net sink for TOC through potential TOC contributions over the last four decades in a tropical region of central–eastern Honduras.

2. Materials and Methods

2.1. Study Area

The study area is located in the municipality of Catacamas, in the central–eastern part of Honduras. This zone has an elevation that varies between 340–359 m.a.s.l., is located between the coordinates 14° 49′ 47″ N and 85° 50′ 40″ W, occupies an area of approximately 313 ha, and is part of the Guayape River valley [19] (Figure 1).

The predominant life zone corresponds to Dry Tropical Forest (bs-T), according to the Holdridge classification [20]. The climate is characterized by an annual rainfall of 1271 ± 209 mm and an average temperature of 25 °C, with minimums of 18 °C and maximums of 35 °C, generating a well-defined double dry-and-wet seasonality [21]. The slope gradient varies from very gently sloping to sloping areas, with slopes ranging from 1 to 10% [22]. The geological formations belong to the sediments group (Qsed) and deposits of siliciclastic sedimentary rocks [23], resulting in the formation of well-developed and very deep soils (>120 cm), classified as Cambisols, with a predominance of loamy to sandy loam textures [24].

2.2. Use and Management of the Different Systems

In the process of selecting and homogenizing the systems studied, aerial photographs and the history of the use and management of these areas were used (Figure 1), identifying and differentiating 3 systems divided into 6 categories (

Table 1. Use and management of the deforested area of the Dry Tropical Forest.

System	Category	Use/Management
Agricultural	Horticultural and Grains (HG)	Deforested in 1982, Zea mays and Oryza sativa were planted under rainfed conditions. In 1985, gravity and flood irrigation techniques were introduced. Since 2007, an advanced system has been implemented that includes mechanization, the use of agrochemicals and fertigation to produce vegetables such as Cucumis melo, Citrullus lanatus, Capsicum annuum, Solanum lycopersicum, Brassica oleracea, etc., in rotation with Z. mays and Phaseolus vulgaris.
	Grains and Cereals (GC)	Deforested in 1982, it was transformed into an intensively rainfed agricultural system consisting of Z. mays and O. sativa. In 2002, gravity and flood irrigation were introduced, while in some cases, drip irrigation is used. Since 2003, the diversity of cultivated species has been expanded, focusing exclusively on grains and cereals, mechanized preparation and the use of agrochemicals. The cultivated species include O. sativa, Sorghum bicolor L., Glycine max and, especially, Z. mays and P. vulgaris.
Agroforestry	Agrosilvopastoral (ASP)	Deforested in 1982 with the aim of producing rainfed Z. mays. From 1988 to 2011, the area was used as pasture. Since 2012, it has been transformed into an experimental production system that includes different agroforestry systems for sustainable production. Some areas are dedicated to grazing cattle (1.5–2 AU ha−1 year−1) and sheep and goats, as well as being used as cut pastures. Small plots have also been allocated for Z. mays and P. vulgaris production.
	Agroforestry Crops (AFC)	Deforested in 1984 to produce horticultural crops. Subsequently, it became an area for forestry and fruit trees, which it still is today. Among the most important species are Cordia alliodora, Tabebuia ochracea, Swietenia macrophylla, Cedrela odorata, Acacia mangium, Salix alba, and the fruit species Mangifera indica, Psidium guajava, Citrus x sinensis, Persea americana, Anacardium occidentale, Averrhoa carambola, Passiflora edulis, Carica papaya, Cocos nucifera and Zea mays. Also, native grasses: Paspalum fasciculatum, Ixophorus unisetus (J.Presl) Schltdl., and Cynodon dactylon.
Pastures	Cutting Pastures (CP)	Deforested in 1984 for natural grazing. Since 1996, it has been used to produce grasses for cutting: Brachiaria hybrid cv. (Mulato II and CIAT BR02/1794) and Panicum maximun cv. Mombaza to obtain fodder and produce silage, hay and quality bales for cattle feed, with 6–7 cuts per year.
	Grazing Pastures (GP)	Deforested in 1982 for natural grazing. In some years, it was in a rest period, nad between 1987 and 1988, the natural cover was eliminated. Pastures were established in 1992 for Brachiaria brizantha cv. (Marandú, MG–4, Xaraes MG–5 and Decumbens) and remain today, with a cattle load of 1.5–2 UA ha−1 per year.

).

The study area, the dry tropical forest (DTF), extended over 313 ha. The deforestation process affected approximately ⅔ of this area, which was distributed among pastures (81 ha), agriculture (67 ha), agroforestry (64 ha), and urbanization (50 ha) (Figure 1).

2.3. Experimental Design and Soil Sampling

We used a randomized experimental design to ensure the selection of independent and bias-free samples [25]. Three systems (agricultural, agroforestry and pasture) were evaluated, in addition to DTF, which was used as a control. The systems were grouped into seven sampling categories distributed in 45 plots with dimensions of 100 m × 100 m: 12 for Grains and Cereals (GC); 6 for each of the following uses; Horticulture and Grains (HG), Agrosilvopastoral (ASP), Agroforestry Crops (AFC), Cutting Pastures (CP), Pastures for Grazing (GP), and 3 for DTF (Figure 1).

Samples were collected in the first 0.30 m of each plot, dividing the sampling depth into three levels (0.00–0.10 m (D1); 0.10–0.20 m (D2) and 0.20–0.30 m (D3)) [26]. Within each plot and for each depth level, 12 subsamples were collected and combined into a single sample, resulting in a total of 1620 subsamples in the entire experiment.

2.4. Soil Analysis

To determine the bulk density, undisturbed soil samples were collected according to ISO 11272–2017 using a metal ring. After sampling, excess soil was removed from both the top and bottom of the ring. The bulk density was then determined by oven-drying at 105 ± 2 °C to constant weight [27].

Soil samples were also collected, air-dried, ground and sieved to a 2 mm diameter. The percentage of coarse fragments (CF) was then determined [28].

Soil organic carbon (SOC) was determined by wet combustion according to the Walkley–Black method [29]. The total organic carbon (TOC) stock was calculated as the product of the organic carbon content, bulk density, sampling depth and percentage of coarse fragments, using Equation (1) [30].

TOC (Mg ha⁻¹) = (1 – (% CF/100)) × (% SOC/100) × (BD) × (D) × (f)

(1)

where TOC (Mg C ha⁻¹) = total organic carbon; CF (%) = coarse fragments; SOC (%) = soil organic carbon content; BD (Mg m⁻³) = soil bulk density; D (m) = sampling depth (0.00–0.10 m, 0.10–0.20 m and 0.20–0.30 m); f = conversion to ha (10,000).

2.5. Statistical Analysis

The values obtained for soil organic carbon (SOC), coarse fragments (CF) and total organic carbon (TOC) at all depth levels (D1, D2 and D3) were grouped and averaged into a single range (0.00–0.30 m). The bulk density (BD) was analyzed separately for each depth level, taking into account the interactions between different land uses and depth in each case. Since the data did not meet the assumptions of normality and homogeneity of variances, a non-parametric statistical analysis was performed using the Mann–Whitney U test with SPSS Statistics v.23 [31].

3. Results and Discussion

The complete results are shown in Table S1 (Supplementary Material: Table S1).

3.1. Bulk Density in the Different Uses

Table 2. Bulk density at various depths evaluated in the Dry Tropical Forest and considered uses.

BD (Mg m−3)	Category	N	Median	Min	Max	Quart 1°	Quart 3°	Mean ± Std
Depth 0.00–0.10 m	Grains and Cereals—GC	12	1.33	1.20	1.42	1.23	1.36	1.31 ± 0.08
	Horticulture and Grains—HG	6	1.29	1.24	1.35	1.28	1.35	1.30 ± 0.04
	Agrosilvopastoral—ASP	6	1.33	1.23	1.54	1.28	1.35	1.35 ± 0.11
	Agroforestry Crops—AFC	6	1.25	1.15	1.41	1.20	1.27	1.25 ± 0.09
	Cutting Pastures—CP	6	1.33	1.20	1.42	1.23	1.40	1.32 ± 0.10
	Grazing Pastures—GP	6	1.36	1.31	1.41	1.33	1.38	1.36 ± 0.04
	Dry Tropical Forest—DTF	3	1.15	1.10	1.16	1.13	1.15	1.14 ± 0.03
Depth 0.10–0.20 m	Grains and Cereals—GC	12	1.30	1.25	1.41	1.28	1.36	1.32 ± 0.05
	Horticulture and Grains—HG	6	1.30	1.26	1.52	1.28	1.35	1.34 ± 0.09
	Agrosilvopastoral—ASP	6	1.37	1.28	1.46	1.34	1.41	1.37 ± 0.06
	Agroforestry Crops—AFC	6	1.29	1.23	1.36	1.26	1.32	1.29 ± 0.05
	Cutting Pastures—CP	6	1.37	1.26	1.46	1.28	1.46	1.37 ± 0.10
	Grazing Pastures—GP	6	1.48	1.40	1.52	1.45	1.50	1.47 ± 0.04
	Dry Tropical Forest—DTF	3	1.25	1.20	1.35	1.23	1.30	1.27 ± 0.08
Depth 0.20–0.30 m	Grains and Cereals—GC	12	1.37	1.32	1.56	1.35	1.41	1.39 ± 0.07
	Horticulture and Grains—HG	6	1.36	1.30	1.64	1.32	1.40	1.40 ± 0.13
	Agrosilvopastoral—ASP	6	1.31	1.29	1.43	1.29	1.36	1.33 ± 0.05
	Agroforestry Crops—AFC	6	1.27	1.20	1.42	1.25	1.28	1.28 ± 0.07
	Cutting Pastures—CP	6	1.42	1.30	1.51	1.32	1.48	1.41 ± 0.09
	Grazing Pastures—GP	6	1.40	1.32	1.46	1.39	1.45	1.41 ± 0.05
	Dry Tropical Forest—DTF	3	1.26	1.21	1.30	1.23	1.28	1.26 ± 0.05

Note. BD: bulk density; N; sampling plots; Min: minimum; Max: maximum; Quart: quartile; Mean ± Std: mean with standard deviation.

and Figure 2 show the BD values at the different depths evaluated in the Dry Tropical Forest (DTF) and in each of the different land uses (pastures, agroforestry systems and agricultural areas).

Deforestation and subsequent land-use change have led to an increase in the BD in the first 10 cm of soil, from 1.14 Mg m⁻³ in DTF to 1.36 Mg m⁻³ in pasture grazing (GP). The use of machinery and animal trampling would be responsible for this increase. In all cases, the increase produced is statistically significant (

Table 3. Bulk density at the 0.00–0.10 m depth in the Dry Tropical Forest and in the different uses.

Category	GC	HG	ASP	AFP	CP	GP
GC	-
HG	ns	-
ASP	ns	ns	-
AFP	ns	ns	ns	-
CP	ns	ns	ns	ns	-
GP	ns	*	ns	*	ns	-
DTF	**	*	*	*	*	*

Note. *: significance level at 5% (p ≤ 0.05); **: significance level at 1% (p ≤ 0.01); ns: not significant (p > 0.05).

), although it is not considered that a process of degradation by soil compaction is taking place, but there is no doubt that, if the current management is maintained, it could occur in the short/medium term.

Something similar occurs in D2, where higher BDs are observed in all uses with respect to D1, even in DTF, although it is not considered that the soils are compacted, since the highest BD reached is 1.47 Mg m⁻³ in GP, probably due to animal trampling. In the other uses, the increase is very small; in fact, significant differences (

Table 4. Bulk density at the 0.10–0.20 m depth in the Dry Tropical Forest and in the different uses.

Category	GC	HG	ASP	AFP	CP	GP
GC	-
HG	ns	-
ASP	ns	ns	-
AFP	ns	ns	*	-
CP	ns	ns	ns	ns	-
GP	**	*	*	**	*	-
DTF	ns	ns	ns	ns	ns	*

Note. *: significance level at 5% (p ≤ 0.05); **: significance level at 1% (p ≤ 0.01); ns: not significant (p > 0.05).

) with respect to DTF are estimated only for appreciation in the GP use and not for the others.

Finally, in D3, the BD values are maintained or experience a slight increase, with the highest BD values corresponding to GP, CP and HG. In all cases, there were significant differences (

Table 5. Bulk density at the 0.20–0.30 m depth in the Dry Tropical Forest and in the different uses.

Category	GC	HG	ASP	AFP	CP	GP
GC	-
HG	ns	-
ASP	ns	ns	-
AFP	**	*	*	-
CP	ns	ns	ns	*	-
GP	ns	ns	*	*	ns	-
DTF	**	*	ns	ns	*	*

Note. *: significance level at 5% (p ≤ 0.05); **: significance level at 1% (p ≤ 0.01); ns: not significant (p > 0.05).

) with respect to the BD of the DTF. The trampling of the animals and the use of machinery, as we have already mentioned in the case of D1, would be responsible for this increase in the BD. However, as we have already pointed out, in no case can we consider the development of a degradation process.

Significant differences were observed at greater depths in GC, HG and GP with respect to the behavior of the BD at different depths in each application (

Table 6. Interaction of the bulk density at different depths within the same use.

Interaction	GC	HG	ASP	AFP	CP	GP	DTF
D1–D2	ns	ns	ns	ns	ns	**	*
D1–D3	*	*	ns	ns	ns	ns	*
D2–D3	*	ns	ns	ns	ns	ns	ns

Note. D1: 0.00–0.10 m; D2: 0.10–0.20 m; D3: 0.20–0.30 m; *: significance level at 5% (p ≤ 0.05); **: significance level at 1% (p ≤ 0.01); ns: not significant (p > 0.05).

).

In general, it is observed that deforestation and subsequent land-use changes cause significant increases in the BD. Allbrook [32] argues that, even in natural areas without human intervention, BD increases with depth, a circumstance we were able to verify in the DTF, as BD experiences a statistically significant increase between D1 and D2, although it no longer differs between D2 and D3. However, activities such as continuous grazing and machinery traffic can increase the BD, as we have described for agricultural and grazing uses.

Decaëns et al. [33] have pointed out that old pastures tend to have higher BD compared with recent pastures or forest areas and that the intensification of management increases the risk of physical degradation of tropical soils. Some authors [34] point out that in agricultural areas where machinery is used intensively, tillage can improve soil aeration, which would result in a lower BD, but this appreciation is not described in our case, since in the agricultural uses studied (GC and HG), BD increases with depth, and this increase is statistically significant with respect to the reference BD (DTF).

Changes in BD reflect variations in the soil structure; therefore, the differences observed in BD between the different land uses are the result of factors related to the management of each one of them. According to Pla-Sentis [35], a minimum of 5 years of tillage is required to observe significant changes in soil physical properties, including BD. In our study, deforestation and the establishment of new uses have occurred in the last 35–40 years.

The presence of compaction, due to an increase in BD, occurs when a value of 1.55 Mg m⁻³ or 1.65 Mg m⁻³ is exceeded in loamy soils or sandy soils, respectively [36]. In our case, this value was reached only in one HG sample in D3 of sandy loam texture, which may indicate the possibility of plow sole development. This circumstance calls for an adjustment in the management of the uses in which machinery is used, since if BD values compatible with the consideration of compacted soil are reached, desertification processes could develop, which would lead to serious environmental and socio-economic consequences.

However, the BD values in the different land-use systems show increasing variations compared with the forest, indicating that compaction is occurring. On the other hand, processes of loss of structure and surface crusting can be observed in agricultural uses (GC, HG). According to Streck et al. [37], these processes are closely related to the indiscriminate and excessive use of tillage tools.

In this sense, variations in this parameter are highly dependent on soil use and management, as well as on the time of management under the same use. These observed changes reflect the interaction between use, management and soil characteristics, highlighting the importance of applying management that is aimed at maintaining or improving soil quality in order to avoid soil degradation and the possible development of desertification processes. Before deforestation, the soils of agricultural, pasture and agroforestry systems were similar and had the same vegetation as the forest.

3.2. Total Organic Carbon Stock in the Different Uses

With the exception of a slight increase in GP, DTF deforestation led to a decrease in TOC stocks for all uses (

Table 7. Total organic carbon stock (Mg ha⁻¹) at 0.00–0.30 m for the Dry Tropical Forest and the different uses.

TOC (Mg ha−1)	Category	N	Median	Min	Max	Quart 1°	Quart 3°	Mean ± Std
	Grains and Cereals—GC	12	68.35	57.27	90.02	65.71	71.29	69.83 ± 8.95
	Horticulture and Grains—HG	6	75.47	45.71	78.90	72.71	77.67	70.94 ± 12.62
	Agrosilvopastoral—ASP	6	75.24	69.34	83.57	70.80	81.33	76.01 ± 6.23
0.00–0.30 m	Agroforestry Crops—AFC	6	44.37	34.71	56.74	38.31	47.74	44.22 ± 8.13
	Cutting Pastures—CP	6	73.73	52.51	113.96	53.15	98.90	77.92 ± 28.04
	Grazing Pastures—GP	6	94.42	54.23	110.42	70.02	108.03	87.99 ± 24.01
	Dry Tropical Forest—DTF	3	86.25	78.51	89.80	82.38	88.03	84.85 ± 5.77

Note. N: sampling plots; Min: minimum; Max: maximum; Quart: quartile; Mean ± Std: mean with standard deviation.

, Figure 3).

The largest losses compared with the DTF stock were observed in AFC, which was reduced by half. In the remaining uses, the losses were not so high. Pastures (GP and CP) are the uses that have practically not changed the stock after deforestation; a similar fact occurs in the use of ASP. In fact, agroforestry uses show a very different behavior.

Paustian et al. [38] point out that tillage alters the structure of the surface layer of the soil and interrupts the continuity of the pore space, thus reducing the soil organic carbon, a situation that is reflected in the use of GC and HG, whose management includes tillage with intensive use of plows and harrows. In addition, the climatic conditions of the tropics favor the mineralization of organic matter [39]. This effect has already been observed in long-term studies, even with the continuous incorporation of high levels of organic residues [40].

The TOC stock in agricultural uses is mainly due to fertilization and the incorporation of crop residues, although there is also the possibility, as indicated by Poeplau et al. [41], that the values in these uses are related to the stabilization of soil organic carbon. In this sense, the authors of [42] point out that the conversion of native forests to cropland can reduce the TOC stock by about 40%. In our case, after more than 40 years of deforestation, the reduction is not so pronounced, representing only 20% in GC and HG, but it is significant (

Table 8. Total organic carbon stock for the Dry Tropical Forest and the different uses.

Category	GC	HG	ASP	AFP	CP	GP
GC	-
HG	ns	-
ASP	ns	ns	-
AFC	**	**	**	-
CP	ns	ns	ns	*	-
GP	ns	ns	ns	**	ns	-
DTF	*	*	ns	*	ns	ns

Note. *: significance level at 5% (p ≤ 0.05); **: significance level at 1% (p ≤ 0.01); ns: not significant (p > 0.05).

). The time elapsed is sufficient for the renewal of the soil organic carbon pools and their stabilization, as shown by several studies [43].

The deforested areas converted into pastures maintain the TOC stock with respect to the DTF, being slightly higher in GP and lower in CP (Table 7). However, no significant differences were observed in any cases (Table 8). As indicated by some authors [44], pastures in tropical regions managed with regulated grazing systems and practices such as the use of improved species, excreta and the presence of live fences and trees contribute to organic carbon storage. In turn, the root system of tropical grasses contributes to maintaining a constant TOC stock, since approximately 80% of the root biomass of B. brizantha is found in the upper 30 cm of the soil [45]. In addition, the absence of tillage favors the formation of microaggregates, where the TOC is protected from microbial action [46]. Dos-Santos et al. [47], in their study of brizantha pastures, have also pointed out the ability of pastures to maintain similar or higher levels of TOC stocks after deforestation, as have Houghton et al. [48], who indicate low losses in the conversion of forests to pastures. Our results, as well as those of other studies, show no significant differences in TOC stocks between the DTF and pastures.

Similarly, pastures in tropical regions managed with regulated grazing systems and practices, such as the use of improved species with C4 metabolism, the contribution of animal excreta and urine (30 to 70% of which is returned to the pasture) and the presence of living fences and trees, contribute significantly to carbon storage [44]. In addition, these effects can also be attributed to the succession process that occurs during the change in use from forest to pasture. In this process, the root system of grasses such as brachiaria plays a fundamental role in the incorporation of organic matter into the soil, helping to maintain constant carbon levels, as observed in the CP and GP uses.

Agroforestry uses (ASP and AFC) show a very different behavior. While the TOC stock is slightly reduced in ASP—in fact, there is no significant difference with respect to DTF—the opposite occurs in AFC, where the largest decrease in stock, about 50%, is detected. The lack of management and the increased susceptibility to erosion due to its location could be responsible for this situation. It is precisely this use that suffers the greatest loss of TOC, being the most sensitive to desertification processes; in this situation, it would be necessary to adopt management measures to stop the deterioration to which it is exposed.

No significant effects on carbon content were observed in unmanaged SAF. In these systems, most of the roots are found in the surface layers of the soil; however, due to the presence of trees and the low surface cover, a significant proportion of these roots reach deeper layers (0.30–1.00 m), reducing the TOC in the surface layers. Our results suggest that the systems evaluated in the SAF, mainly the AFC, could have low rates of organic matter decomposition and, therefore, an increase in carbon storage, perhaps due to a decrease in biological activity.

In general, there is no clear pattern in the scientific literature that consistently shows a significant difference in carbon accumulation between forests and grasslands, although differences are observed when compared with other land uses, as observed in our study. This suggests that the relationship between carbon accumulation and land-cover type may vary considerably depending on environmental, historical and management factors.

When comparing the TOC stock between uses (Table 8), it is observed that there are no differences between agricultural uses (GC and HG), nor between pastures (CP and GP), but there are differences between agroforestry uses (ASP and AFC).

De-Azevedo [49] showed that in the Amazon biome, agricultural areas had 46% less TOC than pastures at a depth of 0.30 m, highlighting the importance of pastures as a fundamental strategy for managing soil organic carbon levels, promoting GHG mitigation and maintaining the sustainability of systems. In our case, when comparing TOC stocks between GP and GC, it was 26% lower in agricultural uses, although this decrease is not significant.

Studies conducted in tropical areas report highly variable TOC stocks, both in forests and in uses and management, similar to those we studied.

For example, in secondary forests in Kenya, TOC stocks range from 174 to 218 Mg ha⁻¹ [50]. In the tropical forests of India [51,52,53], values between 24 and 82 Mg ha⁻¹ are reported. Srinivasarao et al. [54] reported 62 Mg ha⁻¹ for the agricultural use of grains and cereals. In Brazil, De-Azevedo et al. [49] report TOC stocks of between 77 and 138 Mg ha⁻¹ in forests, between 68 and 145 Mg ha⁻¹ in pastures and between 36 and 64 Mg ha⁻¹ in agricultural systems [55], while in sub-humid tropical zones, they report TOC stocks of around 25 Mg ha⁻¹ in agroforestry systems.

The TOC stocks in the first 0.30 m of soil in our study area (313 ha) are 19,624 Mg: 4367 Mg in DTF; 6840 Mg in pastures; 4671 Mg in agricultural uses and 3746 Mg in agroforestry systems. The TOC stocks in our study area, prior to deforestation, are estimated at 26,555 Mg, and the transformation into agricultural, agroforestry, pasture systems and, of course, urbanization, has resulted in a 44% reduction in TOC stocks, which shows how the anthropogenic activities carried out have resulted in a significant loss of biomass and, consequently, a decrease in carbon sequestration, as well as an increase in the risk of loss of organic carbon stored in the soil, as it is more exposed to the action of erosive processes.

In summary, variations in BD are highly dependent on land use and management, as well as on the time of management for the same use. These variations highlight the need to apply management practices aimed at maintaining and/or improving soil quality to avoid the deterioration of its physical properties and the possible development of desertification processes.

Deforestation, which affects forests in general and tropical forests in particular, reduces their function as carbon sinks. In our study, TOC stocks varied from −17% in agricultural areas to −48% in agroforestry systems, making these areas net carbon emitters, except for pastures, which behaved as sinks.

Although pastures had high carbon contents compared with the other land uses, this did not imply a better system status, since it is precisely in these areas that the highest BD values are found.

In this sense, estimates of total organic carbon stocks indicate how human activities can lead to the degradation of forest areas, resulting in a significant loss of biomass and, consequently, a proportional reduction in the carbon stored in the soil. These anthropogenic disturbances are increasing due to population growth and the pressure on natural resources, exacerbating the burden on livelihoods.

4. Conclusions

After 35–40 years of deforestation in a Dry Tropical Forest and the dedication of this area to agro-livestock and agroforestry uses with a high use of agrochemicals and high mechanization, the soil bulk density has experienced a significant increase at all depth levels analyzed, without causing soil compaction. However, it is not excluded that in the short or medium term, this consideration will be reached in both agricultural and livestock uses.

The distribution of soil organic carbon was heterogeneous and sensitive to use and management, so knowledge of its dynamics and temporal distribution can contribute to improving monitoring systems in areas under anthropogenic influence, since soil carbon content is a good indicator of environmental changes.

Both agricultural use and agroforestry showed losses in total organic carbon stocks of the order of −17% and −48%, respectively. However, pasture did not show significant differences with respect to the Dry Tropical Forest, suggesting that proper pasture management can contribute to maintaining the carbon sink potential of deforested tropical areas.

Deforestation and subsequent conversion to agricultural land are the main causes of land degradation, which, in a climate change scenario, may lead to the development of desertification processes, especially in critical and sensitive biomes such as the Dry Tropical Forest.

In future studies, we plan to evaluate the effect of land-use changes on the carbon stored in plants, with the aim of analyzing carbon turnover and storage in natural versus anthropogenic systems, which can contribute to the development of forest and rangeland management models in tropical environments.