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Review

Comparison of Chemical Soil Properties of Temperate Grassland and Arable Land—A Review

by
Matthias Filipiak
1,2,* and
Katrin Kuka
1
1
Institute for Crop and Soil Science, Julius Kühn Institute (JKI)—Federal Research Centre for Cultivated Plants, Bundesallee 58, 38116 Braunschweig, Germany
2
Thuenen Institute of Agricultural Technology, Bundesallee 47, 38116 Braunschweig, Germany
*
Author to whom correspondence should be addressed.
Soil Syst. 2026, 10(1), 20; https://doi.org/10.3390/soilsystems10010020
Submission received: 20 November 2025 / Revised: 12 January 2026 / Accepted: 19 January 2026 / Published: 22 January 2026
(This article belongs to the Topic Carbon and Nitrogen Cycling in Agro-Ecosystems and Other Anthropogenically Maintained Ecosystems—2nd Edition)
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Abstract

Chemical soil properties contribute to the resilience of soil ecosystems. Healthy soils with optimal nutrient levels, balanced pH and good organic matter content are better able to withstand environmental stresses, such as drought, disease or pests. When comparing the chemical soil properties of temperate grassland and arable land, several differences can be observed due to differences in soil cover and management. Grasslands typically sequester more carbon, limit nitrogen leaching, and have lower nitrous oxide emissions and losses of phosphorus due to less soil disturbance and a more closed nutrient cycle. In contrast, arable land has higher nutrient losses through harvest, leaching, gaseous emissions and erosion due to regular tillage, frequent bare phases, and sequesters less carbon, typically due to higher mineralisation rates and lower nutrient returns. Monitoring and managing chemical soil properties, appropriate nutrient management, addition of organic matter such as organic fertilisers, inclusion of grassland phases and catch crops in crop rotations, incorporation of crop residues into the topsoil after harvest and further sustainable agricultural practices are essential to promote soil health. By optimising chemical soil properties, farmers and land managers can improve productivity, conserve natural resources and support the long-term sustainability of the soil ecosystem.

1. Introduction

In addition to biological and physical properties, good chemical conditions are essential for healthy soils [1,2]. The chemical properties and ecosystem services that can be provided by the soil depend on site conditions, such as climate and parent material, and particularly how it is managed to produce food, fodder or renewable biomass [3,4,5]. Ecosystem services related to the chemical status of the soil are nutrient cycling, carbon (C) sequestration and thus climate regulation [2,4,5]. Conserving and managing soil ecosystems in a sustainable way is crucial to maintain these services and ensure long-term environmental sustainability [2,5,6,7]. If the soil is not managed sustainably, there is a risk of, e.g., increased greenhouse gas (GHG) emissions, nutrient leaching, heavy metal contamination, decline in organic matter (OM), acidification or salinisation [8,9,10,11,12].
In general, grasslands provide more non-productive ecosystem services than arable land; farmers do not directly benefit from these services, although they are essential for maintaining soil quality, biodiversity and regulating nutrient and water cycles [4,13,14,15,16]. Extensive grasslands provide more regulation than production ecosystem services, although intensive grasslands with a higher focus on production ecosystem services are necessary to ensure high fodder quality [17,18,19,20,21,22]. Arable land fulfils the primary role of producing food and energy crops, although several practices also aim to improve the provision of ecosystem services [23,24,25,26,27]. However, especially high-productivity systems on both arable land and grassland under suboptimal management (e.g., frequent tillage, heavy field traffic, overstocking, residue removal) can have negative environmental impacts such as loss of soil organic carbon (SOC), GHG emissions, nitrogen (N) leaching or losses of phosphorus (P) and off-site eutrophication [27,28,29,30,31,32,33,34,35,36,37,38]. Since the demand for animal products is expected to grow by 60 to 70% by 2050 [39], current C accumulation rates are insufficient to accomplish sustainable and climate-neutral agriculture considering the GHG emissions from arable soils and livestock [40,41,42,43,44,45]. Reducing the GHG emissions and nutrient losses, and increasing C sequestration, are therefore key objectives of current agro-environmental policies [46,47,48,49].
This literature review aims to provide a scientifically grounded basis for decision-making by farmers, advisors, and policy makers. By comparing key chemical soil properties of mineral soils under temperate grassland and arable land, it identifies management-related conditions that favour C sequestration, reduced N leaching and gaseous N emissions, lower P losses, and mitigation of soil acidification. The synthesis of evidence highlights how land-use-specific practices can be leveraged to maintain soil fertility and enhance resilience to climate change. In this way, the review translates differences in chemical soil properties into actionable insights for sustainable land management and agri-environmental policy design. The primary areas of interest are temperate grasslands and arable land in Europe. In order to compare the effects of land use on soil properties, this review will mainly focus on the topsoil (upper 0–30 cm), while also considering transportation processes into the subsoil and groundwater [50,51,52]. The topsoil layer thereby has the highest plant root density, soil organic matter (SOM) and nutrient content, and is the main zone of management effects [50,53,54,55,56].
For each section, this review provides a tabular summary of the results presented in the cited literature. The comparison of a given property between grassland and arable land is simplified into the categories of being significantly larger on grassland, greater on arable land or not significantly different. The tables also include details about the sampling strategy, number of sites, study region, and sampled depth/horizon. Soils were identified according to the world reference base for soil resources (WRB) [57]. Figure 1 also provides a simplified visualisation of the inputs, outputs and transformation processes of C and N, which are the main nutrient pools in agricultural soils.

2. Results and Synthesis: Chemical Soil Properties

2.1. Comparative Patterns of Soil Organic Carbon Under Grassland and Arable Land

The soil C pool is the third-largest of all global C pools and interacts with the oceanic, biotic, geological and atmospheric pools through various transport and transformation processes [71]. Many important regulating ecosystem services are associated with the SOC content, such as water filtration and storage, maintenance and development of the soil structure, nutrient availability and storage, regulation of pollutants and buffering of the soil pH [72,73,74,75,76,77]. The SOC content is determined by the geological parent material, soil formation processes, climatic conditions and management [78]. Management determines the C input/output ratio and affects the physical, chemical and biological soil properties [30,32,40,79,80]. This leads to changes in the turnover conditions that affect the SOC cycle, resulting in either C emissions or sequestration [81]. A major part of SOC is located in the topsoil, where it is at risk of being lost due to erosion, decomposition and disturbance [51,67]. Various processes can transport SOC from the topsoil to the subsoil, where it is less vulnerable to disturbance and can be stored for the long term [40,59]. A summary of literature comparing SOC-related differences between arable land and grassland as an effect of land use can be found in Table 1.

2.1.1. Processes Mediating Changes in Carbon Stocks

Understanding the processes that cause changes in the SOC turnover allows for a long-term implementation of management practices that foster SOC sequestration. The key factors influencing the long-term storage of C in the terrestrial pool are (1) the available specific surface of mineral particles, (2) input and output rates of C, (3) biochemical decomposability of SOC and (4) physical protection of aggregates [40].
Specific Surface of Mineral Particles
The available specific surface of mineral particles is a limiting factor for SOC storage [78,108] due to a significant and positive correlation between SOC on the one hand and the clay and silt content on the other [81]. Thus, in coarse soils with low mineral surface, differences in SOC accumulation caused by management may be smaller than on sites with a higher specific mineral surface [109]. Accordingly, evidence was presented indicating that 86 to 91% of SOC was associated with the mineral fraction of silt and clay in both grassland and arable soils [110].
Input and Output Rates of SOC
Given sufficient available mineral surface, a positive SOC balance is achieved when C inputs, e.g., from belowground biomass production, retention of litter (including harvest residues), and organic fertilisers application, exceed losses, e.g., by harvest, emissions or erosion. Perennial grassland vegetation is characterised by a dense root network, continuous belowground biomass production and higher root C contents than arable land; in contrast, arable land often has a sparser root system due to its annual crop rotation and regular tillage, resulting in reduced belowground SOC allocation [53,58,61,74,82,87,101,111,112,113]. Consequently, increasing the mowing frequency on grasslands promotes SOC sequestration as an effect of vegetation and root regrowth [114,115,116]. In contrast, arable land is at higher risk of C losses, especially if all crop residues are removed [59,65,101,117]. Grasslands also typically receive higher application rates of C-rich organic fertilisers, whereas arable land typically receives C-deficient mineral fertilisers [63,118,119]. While SOC sequestration benefits from inputs of N by increased plant growth and can be limited if N is deficient [63,77,111,120,121,122], on European soils N inputs are often higher than the requirements [30]. The substitution of mineral fertilisers with organic fertilisers is therefore a viable method to meet the N demand and increase the C inputs on arable land [63,80,123]. The fertilisation regime was in fact reported to be the primary reason for differences in SOC contents under different arable crop types [124]. The inclusion of forage legumes, such as alfalfa, clover, and vetch, into arable crop rotations can also substantially promote SOC sequestration [125]. While the losses in the form of CO2 emissions are higher on grassland due to higher microbial activity, the higher inputs and returns result in an overall positive net SOC balance [58,69,83].
Biochemical Decomposability of SOC
The chemical decomposition of SOC is the result of biological metabolism by the activity of soil biota, which is controlled by aeration, temperature and soil moisture [74,81,101,126,127]. The transformation of labile OM fractions into more stable fractions is fostered under undisturbed conditions [104], while regular tillage activities increase aeration and mineralisation, and reduce the stabilisation of labile SOC in minerals [16,74,103]. A higher proportion of stable SOC is thus usually found in grassland soils or arable soils under no or conservation tillage [103,104,105,106].
Stabilisation in Soil Aggregates
Since the majority of SOC is stored in macroaggregates, the aggregate stability is directly related to a site’s potential to sequester SOC [40,128,129]. Soil aggregates and SOC form a reciprocal relationship, since on the one hand SOC stabilises and forms aggregates due to its binding properties, while on the other hand aggregates protect therein stored SOC from decomposition [40,72,73,97,110,128,130]. On grassland soils a high abundance of perennial roots and fungal hyphae further stabilises aggregates [61,87,102,131,132,133], while on arable land regular tillage causes physical aggregate breakdown and exposes the intra-aggregate surfaces containing SOC to decomposers [40,72,97,101,110]. Furthermore, the intra-aggregate protection of SOC may also be limited by management practices that result in soil compaction and thus hinder root growth [132,134], e.g., the use of heavy machinery, uncontrolled and intensive grazing, and field traffic during conditions when the soil is particularly susceptible to compaction [30,31,32].

2.1.2. Vertical Distribution and Incorporation of SOC into the Subsoil

The majority of sequestration effects occur in the topsoil [51,59,67,120]. However, the long-term SOC storage is fostered by the incorporation of SOC into deeper layers, where slower turnover processes occur due to lower oxygen concentrations and fewer disturbances [50,78,101,135,136]. Due to fundamental differences in management, land cover and translocation and mixing processes, arable and grassland exhibit different vertical SOC distributions across the soil profile [15,31,51,56,78,101,137]. The key processes affecting the vertical distribution of SOC are (1) tillage and (2) bioturbation, while translocation and losses as dissolved organic C are generally considered as minor in both arable and grassland systems [78,83,138].
Tillage
Tilling of arable land has a homogenising effect on SOC contents in the tillage layer (typically the top 30 cm), whereas grasslands and non-tilled arable land show a gradual decline in SOC with increasing depth [31,53,58,65,78,139]. The conversion of grassland to tilled arable land typically results in SOC changes in the tillage layer, while lower layers are usually unaffected [81,90]. A less common, typically one-time, tillage practice applied on arable land is full inversion or deep tillage: with the mechanical inversion of the profile, the SOC-rich topsoil is buried beneath the usual tillage layer and subsoil substrate with low SOC content is brought to the surface, where sequestration can resume [140,141,142,143,144].
Bioturbation
Bioturbation is the process of mixing soil layers by burrowing activities of soil biota, especially deep-burrowing earthworms [83]. The permanent vegetation of grassland, particularly grazed and extensive, provides favourable conditions for a high diversity and abundance of soil fauna, thereby promoting the transport of SOC into deeper layers by bioturbation [78,83,84,100,133,145,146,147,148,149,150]. In contrast, regular tillage and seedbed preparation on arable land reduce earthworm abundance due to physical disturbance, unless under no-tillage management [83,84,145,151].

2.1.3. Response of Land Use Change on C Balance and SOC Stocks

Former land use can affect the soil properties up to a century after conversion [63,152,153] and consequences of changes in management and land use on SOC are often not apparent immediately, but after several years to decades [42,130,153,154]. Low N availability [15] or too extensive management [62] are potential causes of slowed SOC sequestration after conversion from arable land to grassland. The absence of tillage during arable use after conversion from grassland may also result in no changes in the short term [82] and high SOC levels on arable land may be a result from residual roots from a previous grassland vegetation [87]. Therefore, in order to understand long-term changes in SOC contents, first the dynamic equilibrium of C, i.e., the time until SOC changes come to a halt after a land use change, needs to be addressed. During the time between land use change until an equilibrium is reached, the soil can act as either a sink or source of C [40,79,96]. Then, the absolute effect of land use on SOC can be discussed, including the effect of different management strategies.
Dynamic Equilibrium
Equilibrium of SOC is reached when the SOC fraction that is lost is equal to the newly formed SOC [96]. After a disturbance or land use change, a new equilibrium can be reached in response to new management practices [40,79]. SOC sequestration rates thereby begin to slow down as they approach saturation [30,92], whereby most SOC is sequestered during the first 20 years [56,92,139,155,156]. In general, an equilibrium is reached faster following conversion of grassland to arable land (few years to decades) than vice versa (up to a century), because the establishment of a perennial vegetation, recovery of the rhizosphere and subsequent SOC sequestration are slower processes than disturbance by tillage and decomposition of SOC [30,58,81,84,95,139].
SOC Content at Equilibrium
While multiple factors influence the potential for sequestration, under otherwise similar site conditions and conventional management, SOC content typically decreases after conversion of grassland to arable land until they reach equilibrium [99,153]. Losses of SOC in the topsoil of arable land after conversion from grassland have been reported to be as high as relative 29 [94], 38 [86] and 59% [79]. Conversely, the SOC content increases when arable land is converted to grassland [30,31,50,55,69,70,85,87,88,89,90,91,93,96,101,102,157]. As a result of long-term accumulation, SOC stocks in grassland can be two [53,99] to three times higher [100] than in arable land.

2.2. Comparative Nitrogen Dynamics Under Grassland and Arable Land

Soil N is an essential plant nutrient that directly affects plant productivity [15] and limits plant growth if lacking, even under optimal water availability [61,158]. In most agricultural systems, the N input is dominated by application of organic and mineral fertilisers, and returns by animal excreta [34,124,158,159]. Losses largely occur by plant uptake, removal by harvest, grazing, leaching and gaseous emissions [60,61,158,160,161,162]. If annual fertiliser application and residual N content in the soil surpass plant uptakes, the resulting N surplus may pollute water and land ecosystems, affect air and soil quality, climate, as well as biodiversity and human health [60,61,63,160,163]. Through N inputs and outputs and its ability to alter turnover conditions in the soil, land use has a profound impact on the ability of soils to store N, provide available N for plants, the N leaching potential, and the gaseous N emissions [34,36,60,62,73,164,165]. Table 2 provides a summary of the effect of land use on total N, available N, N leaching and gaseous N emissions.

2.2.1. Land-Use-Related Changes in Nitrogen Content

Changes in the soil N stocks must be considered in terms of both (1) total soil N, which consists largely of organically bound N, i.e., immobilised and stored in SOM and microbial biomass, and (2) exclusively mineral N [61,175,176]. Following SOM mineralisation, mineral N forms, primarily nitrate (NO3−) and ammonium (NH4+), are released into soil solution and are available for plant uptake, although they may also be subject to various mechanisms of losses [36,60,61,62].
Total N Stocks
Grasslands typically have a higher total N storage capacity and a higher potential to be N sinks than arable lands [34,55] due to their higher content of SOC and its ability to store N [73,176]. The effects of land management on total N content in soils largely prevail over site-specific effects, such as climate, topology and lithology [83,98,166,177]. Thus, higher total N content in grassland than in arable land was reported in unfertilised pastures due to higher SOC content [164] and losses in total N after conversion of grassland to arable were found to account for 61% in the upper 15 cm [53]. Total N stocks may be increased in the long term by management practices that foster SOC sequestration [85,102] or by the inclusion of legumes into grassland mixtures and catch crops on arable land [62,157,171,178]. However, the N accumulation slows down with increasing N content, as it approaches saturation [157].
Available N
Compared to total N, available N is more closely related to short-term management effects, especially N fertiliser inputs, and shows pronounced seasonal variations [60,62]. The mineralisation, removal, and risk of losses of available N is higher in tilled soils grown with annual crops, while losses are lower under an undisturbed permanent vegetation due to larger biomass returns, and a continuous N uptake and cycling [58]. Consequently, available N content might not change significantly after conversion of unfertilised grassland to fertilised cropland [81], while elevated available N content due to high fertiliser application during prior arable cultivation might decrease slowly after conversion to grassland [58,62,63]. Variation in soil N between different arable crop types in turn is primarily driven by fertilisation regimes, i.e., amount and type (organic vs. mineral fertiliser) [124,179]. A viable method to at least partly counteract higher losses of available N under arable cultivation may be rotational farming including ley phases, which was shown to result in increased soil fertility, higher yields and reduced N fertiliser requirement over a time span of 31 years compared to permanent arable land [180]. However, the intervals of rotational farming required for the manifestation of effects related to vegetation cover may be as long as 3 years due to the time needed to establish a new root system [58,62].

2.2.2. Nitrogen Leaching Potential

Leaching of N is the process of mobilisation of soluble N and subsequent seepage into groundwater, where it poses a threat to health and the environment. A major part of the leachable N fraction comprises NO3−, while leaching as NH4+ can be considered negligible due to its high affinity to be adsorbed by the surface of clay minerals and typically low NH¬4+ concentrations due to a rapid conversion to NO3− [58,60,181]. The management-related factors that control N leaching are (1) the balance between N input (e.g., from fertilisers or animal excreta) and crop uptake, (2) soil water availability and permeability, (3) N immobilisation (i.e., by microbial biomass or adsorption by SOM) and (4) plant cover and diversity [52,60,165,171,182].
N Input and Uptake
The balance between N input and plant uptake determines the N surplus and thus has a direct impact on the risk of leaching. Compared to arable land, grasslands have a higher and more constant N uptake from the soil due to a longer growing season and perennial vegetation [167,183,184]. This results in a lower N leaching risk on grassland, even in cases of higher N fertilisation rates [35,60,185]. On arable land, the leaching risk is particularly high during the bare periods between harvest and crop emergence due to lacking uptake [168,186]. This risk can be alleviated by reducing the total amount of fertiliser applied, avoiding fertiliser application during periods with low expected uptake, planting catch crops to maintain a constant N uptake or by an implementation of fodder grass within a crop rotation [35,36,60,187,188].
Soil Water Availability and Permeability
Leaching of N depends not only on the N availability, but also on the amount of water available for infiltration and percolation [52,189,190,191,192,193]. Differences in vegetation cover (perennial vs. annual) thereby explain the large differences in leaching losses via the impact on evaporation and thus soil moisture [58,168,188]. In the absence of catch crops, the risk of N leaching is typically highest on arable land in autumn and winter, when the soil moisture is highest, during the bare phases after tillage and before crop emergence; it reaches a minimum in summer, when water uptake increases and the soil moisture decreases [175,178,187,194,195]. In addition, tillage both increases infiltration in the short term and has long-term negative effects on soil structure [88,97,196], contributing to the higher leaching risk of arable land [58,164,169].
N Immobilisation
A reduced risk of N leaching can be achieved by increasing N immobilisation [58,164,175,176,185,197]. Particularly in biodiverse grassland soils, the competition for NH4+ among microbiota and nitrifying bacteria is higher than in arable soils, so that a greater proportion of N is consumed and immobilised in microbial biomass and SOM, whereas arable soils have an increased activity of nitrifying bacteria, which increases NO3− availability [58,172]. On arable land, particularly, the retention of crop residues is thus a viable method to reduce soil NO3− concentrations in the post-harvest period through the assimilation of N into SOM [176,197,198].
Plant Diversity
A high biodiversity of grassland flora was shown to contribute to a low leaching risk by reducing the need for fertilisation through a more closed N cycle and constant uptake [37,170], as also evidenced by a positive correlation between plant diversity and available N [171,199].

2.2.3. Gaseous Nitrogen Emissions

Gaseous N emissions occur as diatomic nitrogen (N2), nitrous oxide (N2O), nitric oxide (NO) and ammonia (NH3) [60,162]. Compared to other pathways of N losses, gaseous emissions account for smaller quantities [161,162]. However, N2O, in particular, is a GHG of concern because it has the 265-fold global warming potential of CO2 [15,48,64] and accounts for the majority of agricultural GHG emissions [200]. Key drivers of N2O emissions are N availability, respiration activity, aeration, and soil moisture [36]. When there is sufficient available N, the highest N2O emissions typically occur as a product of anaerobic denitrification with elevated soil moisture, usually in spring or after rewetting of the soil [36,64,159,160,200]. Emissions of N2O are consequently greatly affected by management practices, particularly (1) fertiliser application and (2) tillage [34,161,201].
N Fertilisation
The input of readily available N from fertilisers, whether in mineral or organic form, is a central and direct cause of N2O emissions in both arable land and grassland by exceeding soil denitrification capacity and plant uptake [36,52,64,159,201]. Emissions of N2O thus typically peak shortly after fertilisation events [200]. This effect even prevails across different climates and paedogenic properties, resulting in increasing N2O emission on both grassland and arable land with increasing N input [33,34].
Aeration, Tillage and Land Cover
Tillage practices are an indirect cause of increased gas fluxes by breaking down aggregates, exposing previously protected SOM, loosening the soil, and increasing aeration and mineralisation [58,61,172]. Increased mineralisation of SOM results in higher N2O fluxes due to elevated microbial availability of NO3− for denitrification [202,203]. Regular tillage practices on arable land are a central cause for increased GHG emissions [52,172,203] and were shown to result in higher N2O emissions than under grassland, even at similar fertiliser application rates [52]. However, the effect of tillage varies with time, since highest tillage-related N2O emissions were associated with freeze–thaw cycles, which were particularly pronounced on soils tilled in autumn, which in turn promoted water infiltration [173].

2.3. Land-Use Effects on Soil Phosphorus Under Grassland and Arable Land

Phosphorus (P) is an essential, often limiting, macronutrient [158,163]. Its natural supply largely depends on weathering of the geological parent material, although this accounts for relatively small amounts [63,204]. The main P pathways in agricultural systems are (1) uptake by harvest or grazing, (2) inputs by fertilisation and animal excreta and (3) losses by erosion or leaching [38,63,98,162,163,167]. Table 3 summarises the effects of land use on soil available and total P content.

2.3.1. Uptake by Harvest or Grazing

The P uptake in the form of grazing and harvest typically exceeds the natural supply in agricultural soils [63,158,204,206]. Thereby, grasslands usually have a more closed P cycle than arable land and a smaller proportion of biomass, and thereby P is removed from grassland soils by harvest and grazing, whereas the plant uptake and removal on arable land is often substantially higher than the return from crop residues [61,73,158,206].

2.3.2. Fertilisation and Animal Excreta

As a result of low natural resupply, maintaining sufficient P content in agricultural soils largely depends on additional inputs [38,63]. Due to higher outputs, arable lands typically receive higher P fertiliser input rates than grasslands [158]. Furthermore, fertilisation regimes are typically the causes for variation in P content between different arable crop types [124,179]. Especially on grazed grasslands the returns with animal excreta further reduce the requirement for P inputs [38,163,206,207]. The majority of the P requirement is thereby met by organic fertilisers on grassland and mineral fertilisers on arable land, although arable sites in cattle-rich regions may receive high organic fertiliser inputs as well [63,166,206,208]. Because organic fertiliser inputs are often determined by the N requirements, P inputs in such systems may exceed the requirements [38,209], posing a risk of overfertilisation. Grasslands are thereby typically at lower risk of overfertilisation due to better P retention capacity [61,63,158,206]. As a result of lower losses under grassland, P content may not differ significantly between P-fertilised arable land and unfertilised but grazed grassland [205], while P content declines under mown grassland in the absence of fertiliser or manure inputs, although at a slow rate [15,38,63,207]. Elevated P content of arable land may therefore not change significantly in the short term after conversion to grassland [62] and might persist for up to 50 years [210], although effects of previous P fertilisation lasting for centuries have also been documented [63].

2.3.3. Losses by Leaching and Erosion

When P inputs surpass the requirements, the resulting losses through either leaching or erosion are at risk of causing off-site damages, such as water eutrophication and loss of biodiversity [38,204,211]. Although P availability was reported to be higher in grassland than in arable land, this did not result in increased P loads in groundwater [167], due to a positive correlation between available P and SOC [72,73]. Nevertheless, P losses by leaching are generally considered to be low due to its overall low solubility [204]. Instead, major P losses are a result of tillage-induced erosion due to the strong binding of P in soil aggregates [38,63,167]. This risk is particularly high on arable land, where tillage and therewith associated topsoil erosion are common [30,38,67,72].

2.4. Contrasting pH Regulation in Grassland and Arable Soils

The soil pH value indicates the acidity of a soil solution, which directly affects the mobility of nutrients and toxic elements [86,212]. Due to differences in vegetation composition, optimal pH values are about 0.5 pH units lower on grassland than arable land [213,214]. The pH value of a soil is largely influenced by (1) the mineralogical composition of the parent material, (2) management practices (liming and fertilisation) and (3) the vegetation cover and SOC contents [213]. Table 4 provides a summary of studies comparing the effects of land use on the soil pH.

2.4.1. Mineralogical Composition

The geogenic carbonate content of the parent material and the clay content determine the base level of the soil, thus its ability to bind cations, buffer pH changes and resist acidification [31,78,86,213]. Due to higher production demands and economic pressures, fertile clay soils in the temperate zone are predominantly used for arable cultivation, while grasslands are more often located on less productive soils [13,193,211,213,216,217]. This bias consequently results in significantly lower reported pH levels in grassland soils than in arable soils due to a spatial correlation of the soil clay content and arable soils [86,213].

2.4.2. Fertiliser Application and Liming

The application of acidifying fertilisers, e.g., sulphur fertilisers, ammonium-based fertilisers or urea, is a key contributor to decreasing soil pH in both arable land and grassland [214,218,219,220]. As per a meta-analysis, the application of acidifying N fertilisers resulted in a significant decrease in soil pH by an average of 0.10 units compared to control treatments [126]. Fertiliser-induced soil acidification is particularly detrimental to nutrient availability on already acidic soils [213,214,221] and may even force the adoption of crop types more tolerant of low pH levels [124]. The pH buffering and raising effect of base cations, e.g., by liming, can thereby be used to counteract acidification and achieve favourable pH conditions [91,98,213,214]. However, an analysis of soil inventory in Germany revealed that 65.4% of arable land and 22.8% of grassland received lime application, although a larger proportion of grassland soils (52%) than arable soils (41%) had pH values below the recommended optimum, probably due to the greater economic importance [213].

2.4.3. Plant Cover

Soil pH is highly dependent on the plant cover due to higher inputs of acidifying organo- and humic acids excreted by roots under a perennial vegetation, which results in lower pH levels in grassland soils than in arable soils [31,70,72,73,98,100,110,215]. The lower soil pH of former grassland or a crop rotation with grass can thereby persist for two [15] to three years or longer [100] after conversion to arable land. However, differences in pH levels between grassland and arable land may be small in cases of similar SOC content and under otherwise similar site conditions [78].

3. Summary and Conclusions

This review summarises and discusses the current knowledge on the effects of grassland and arable land use in the temperate zone on chemical soil properties, considering the provision of ecosystem services (e.g., SOC sequestration, mitigation of N leaching and gaseous N emissions, nutrient provision, regulation of soil pH). In addition to the effects related to long-term arable and grassland use, this paper also addresses the effects of short-term changes. Arable and grassland use requires sustainable management practices in order to maintain soil quality as well as nutrient availability to ensure a resilience to climate change, which will remain a challenge for future decades. Nevertheless, this is often a balancing act between several viable options and conflicting interests, as there is no single “silver bullet” solution for achieving both maximal sustainability and productivity on one site. A preferred land use, whether arable land or grassland, should be strictly adapted to the climatic and paedogenetic conditions in situ, taking into account the advantages and disadvantages of the sites under both uses, while opting for reduced management intensity to compensate for negative management-related effects.
Arable land and grassland provide different sets of ecosystem services, with arable land focusing on provisioning services and grassland balancing between provisioning and regulating services. Arable land is essential for the production of food and energy crops, but the negative effects of management on soil quality, high risk of GHG emissions, nutrient losses and associated off-site damage put it in a tight spot between economic and environmental demands, with the former usually taking precedence. In contrast, grassland is more beneficial in terms of soil quality, maintaining biodiversity, SOC sequestration and nutrient loss reduction, although these benefits are often overshadowed by the questionable profitability.
Unless under no or conservation tillage, regular tillage and the associated increased aggregate breakdown, aeration and subsequent biochemical decomposition are central causes of declining SOC under arable land, which also affects long-term fertility. In contrast, perennial and undisturbed vegetation under grassland promotes SOC sequestration. The frequent removal of crop residues on arable land also contributes to lower C returns, whereas on grassland a greater proportion of C is returned to the soil via residues and animal manure. Grasslands also typically receive larger amounts of organic fertilisers with high C content whereas mineral fertilisers with low C content are common on arable land. Nevertheless, higher SOC sequestration under grassland can be counteracted by emissions from grazing cattle, especially at high stocking densities.
The application of mineral fertilisers under arable land leads to both higher NO3− leaching and gaseous N2O losses. The risk of leaching is particularly high immediately after fertilisation and tillage, when the soil is loosened, and during bare periods without vegetation cover, due to lack of uptake and increased moisture typical of arable land in autumn and winter. On the other hand, grassland vegetation, which is particularly rich in biodiversity, limits leaching through constant uptake and high content of SOC, which adsorbs NO3−. Furthermore, regular tillage of arable land prevents a complete reduction in NO3−, with N2O being emitted as a by-product in large quantities. The typically higher mineral N input and lower SOC content under arable land further contribute to higher N2O emissions by increasing the availability of NO3− for denitrification. In particular, thawing of previously frozen soils and increasing soil moisture have been associated with increased N2O emissions. This effect is particularly pronounced on soils that are kept bare in winter due to lack of water uptake, while water fluctuations and the corresponding moisture-driven N2O emissions are lower under perennial vegetation.
Another concern in agricultural soils is the availability and accumulation of P, which is sometimes a limiting nutrient. In particular, arable land has a larger P removal by harvest. In contrast, on grazed grassland, a greater proportion of P is returned to the soil with animal excreta, and the higher SOC content further promotes P storage and reduces losses. Regular tillage practices and bare periods on arable land also contribute to increased P losses through topsoil erosion. The higher total P losses on arable land motivate a higher need for additional P fertilisation. The required P is often supplied in combination with N, especially when applied as organic fertiliser. Since the amount of applied fertiliser is often determined by the N requirement and the P requirement is lower than the provided input, there is a substantial risk of P overfertilisation. Especially when eroded into surface water, high P content is a cause for eutrophication. Elevated P content under grassland is slow to decline due to lower losses and has been reported to persist for decades to centuries, sometimes even preventing the establishment of biodiverse vegetation.
Several effects lead to lower pH values in grasslands than in arable soils: on the geogenic side, grasslands are more often located on acidic parent material, whereas arable cultivation is preferred on clay soils with higher pH values and better pH buffering capacity; on the biogenic side, higher SOC content under permanent grassland vegetation reduce the soil pH; on the management side, due to economic incentives, arable soils are more often treated with lime to counteract soil acidification compared with grassland soils. While a lower pH generally favours nutrient availability under grassland, severe acidification can be detrimental, increasing nutrient losses and mobility of toxic elements.

4. Recommendations

Chemical soil properties can be largely improved by fostering SOC sequestration. High SOC content improves nutrient availability and also limits N leaching, nutrient losses through erosion and gaseous N emissions. Most management practices that improve chemical soil properties are aimed at “closing” the nutrient cycle, i.e., reducing outputs and increasing returns. Based on the findings presented in this review, we recommend that the following management practices should be implemented, if possible:
  • Increase OM inputs on arable land by retaining crop residues, alternating grassland phases within an arable rotation and substituting mineral fertilisers with organic fertilisers;
  • Determine organic fertiliser inputs based on N and P requirement to avoid N and P overfertilisation and supply the remaining requirement for N and P through mineral fertilisers;
  • Include legumes in crop rotations and grassland mixtures to increase biogenic N fixation and reduce N fertiliser requirements;
  • Improve nutrient cycling by fostering a diverse vegetation and avoiding monocultures;
  • Reduce erosion by avoiding bare periods on arable land by planting catch crops and implementing reduced and conservation tillage;
  • Reduce leaching by avoiding tillage during the wet season from autumn to winter, especially when soil mineral N is high;
  • Prefer grazing over mowing on grassland, but also avoid overstocking;
  • Monitoring soil pH, especially if acidifying fertilisers are applied, and liming the soil if necessary to counteract acidification.

Author Contributions

M.F.: Methodology, Investigation, Data Curation, Writing—Original Draft, Writing—Review and Editing, Visualisation. K.K.: Conceptualisation, Writing—Review and Editing, Supervision. All authors have read and agreed to the published version of the manuscript.

Funding

The research was funded by the Federal Ministry of Agriculture, Food and Regional Identity (BMLEH) [grant number 2818300916]. The research project is assigned to the guideline on the funding of innovations for sustainable grassland management in the programme for the promotion of innovation of the Federal Ministry Agriculture, Food and Regional Identity (BMLEH). It was carried out by the Federal Agency for Agriculture and Food (BLE) within the framework of the Innovation Promotion Programme.

Data Availability Statement

No new data were created or analysed in this study. Data sharing is not applicable to this article.

Conflicts of Interest

The authors declare no conflicts of interest.

Abbreviations

The following abbreviations are used in this manuscript:
ALArable Land
CCarbon
CO2Carbon Dioxide
GHGGreenhouse Gases
GLGrassland
OMOrganic Matter
NNitrogen
N2atmospheric Nitrogen
NH3Ammonia
NH4+Ammonium
NONitrous Oxide
NO3−Nitrate
N2ONitrous Oxide
PPhosphorus
SOCSoil Organic Carbon
SOMSoil Organic Matter
WRBWorld Reference Base for Soil Resources

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Soilsystems 10 00020 g001
Figure 1. Simplified soil C and N cycle of agricultural soils, with major inputs (blue) and outputs (red) as well as transformation and transportation processes (black, green and yellow) denoted in cursive, major nutrient pools and transformation products in bold, according to information gathered from [34,36,40,52,58,59,60,61,62,63,64,65,66,67,68,69,70], visualised using BioRender.com (accessed on 21 May 2025).
Figure 1. Simplified soil C and N cycle of agricultural soils, with major inputs (blue) and outputs (red) as well as transformation and transportation processes (black, green and yellow) denoted in cursive, major nutrient pools and transformation products in bold, according to information gathered from [34,36,40,52,58,59,60,61,62,63,64,65,66,67,68,69,70], visualised using BioRender.com (accessed on 21 May 2025).
Soilsystems 10 00020 g001
Table 1. Summary of results measuring the effect of land management on carbon dynamics, in particular contents of soil organic carbon (SOC) and soil organic matter (SOM), stable organic matter (OM), and light fraction OM, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
Table 1. Summary of results measuring the effect of land management on carbon dynamics, in particular contents of soil organic carbon (SOC) and soil organic matter (SOM), stable organic matter (OM), and light fraction OM, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
PropertyRelationNumber of SitesSampling StrategySampling TimeDepth [cm]Soil Type/Texture ClassStudy RegionReference
SOC/SOMGL = AL2paired sites, repeated samplingJune 2006–June 20070–100Cumulic PhaeozemOttawa County, Kansas, USA[82]
6chronosequence2008–20090–30Silt, clay, clay loam, silty loamWellesbourne, England[15] 1
8paired sites, chronosequenceMarch and July 2010 and 20110–10Silty loam, Brown EarthSoutheastern Scotland[62]
12paired sitesn. a. 0–30Gleyic/Haplic Leptosols, Haplic/Stagnic CambisolsOuter Western Carpathians, Poland[83]
GL > AL2paired sitesJuly 2004–May 20050–30 n. a.Northeast Pakistan[73]
4paired sitesn. a.0–5n. a.Central Japan[84]
4paired sites, repeated sampling2005–2008 (36 months)0–10CambisolLusignan, France[58]
1repeated sampling1991–20120–15Dystric CambisolAberdeen, NE Scotland[85]
1repeated samplingn. a.0–20Clay loamIndagai Mountain Pass, Cankiri, Turkey[86]
179paired sites, repeated sampling1978–1988, 2007–20090–100Alluvial soils, Brown Earths, Gleys, Peats, Podsols and RankersScotland[59]
4paired sitesMay 20040–50Stagnic Vertisol, ArenosolThuringia, Germany[78]
2paired sitesMay and June 20080–20n. a.North Central Kansas, USA[87]
2chronosequenceMay 19970–20Sandy loam, loam, clay loamWatkinsville, Georgia, USA[55]
18paired sites20080–30FluvisolKolubara Valley, Western Serbia[88]
n. a.repeated sampling1955–20050–30VaryingSouthern Belgium[89]
74 studiesmeta-analysis (paired sites, chronosequence, repeated sampling)n. a.n. a.n. a.16 countries, focusing on Australia, Brazil, New Zealand, USA[79]
2paired sites, repeated samplingOctober 20080–75Chromic Luvisol (silty clay loam)Rothamstedt, Harpenden, UK[53]
n. a.reviewn. a. n. a.n. a.New Zealand[61]
3repeated sampling1987–20060–40 Orthic Luvisol (sandy loam)Tänikon, Switzerland[90]
3paired sites, repeated sampling2014–20160–30Haplic LuvisolUhřice, Czech Republic[91]
n. a.meta-analysisn. a.n. a.n. a.Europe[42]
273 (65 studies)meta-analysis (paired sites)n. a.0–20Terrestrial mineral soils of the temperate zoneNorthern Hemisphere (Central Europe, North America, Russia)[92]
836 (235 studies)meta-analysis (paired sites)n. a.0–100+n. a.Mostly south and north temperate zone[74]
6paired sites, repeated sampling2010–20170–30Eutric Luvisol/CambisolNorthern Germany[93]
12paired sites, repeated sampling1949–20090–20LuvisolZurich, Switzerland[94]
6paired sitesn. a. 0–30Vertic CambisolCalabria, Italy[31]
24paired sites, chronosequencen. a.0–80VaryingEurope[50]
322 (95 studies)meta-analysis (mostly paired sites)n. a.0–30 (±6)n. a.Temperate zone, Worldwide[95]
2paired sites, repeated sampling1999–20100–20Eutric CambisolCzech Moravian Upland[96]
3paired sitesNovember 19950–20Fine-silty loamSydney, Australia[97]
313chronosequence1995–20010–10n. a.New Zealand[98]
11chronosequencebefore 1995, 1995–2000top horizonSandy (predominant), siltyNorthwest Germany[99] 2
398 (81 studies)meta-analysisn. a.0–100+n. a.Worldwide[81]
4paired sitesOctober 2002–October 20040–10Sandy loamMelle, Belgium[100]
717chronosequence2000 and 2004top horizonCambisols (mostly), Leptosols, Regosols, Stagnosols, Albeluvisols, Planosols, GleysolsBavaria, Germany[70]
717chronosequenceCollected after 1990, main part between 2000 and 20040–100+ Cambisols (mostly), Leptosols, Regosols, Stagnosols, Albeluvisols, Planosols, GleysolsBavaria, Germany[101]
6chronosequenceJuly 20120–10Calcic-Orthic Aridisols (sandy loam)Inner Mongolia, Northern China[102]
Stable OMGL > AL4paired sitesMarch 20156–15 cmChromic Luvisol (silty clay loam)Rothamstedt, Harpenden, UK[103]
8paired sites20120–30Cambisols and FluvisolsKlamputė and Dembava, Central
Lithuania		[104]
6paired sitesn. a.0–30Sandy soilsHoshiarpur, Northwest India[105]
5paired sitesMay and October 20120–30Calcic ChernozemNovi Sad, Serbia[106]
Light fraction OMGL > AL154 (28 Studies)meta-analysisn. a.n. a.n. a.n. a. [107]
5paired sitesMay and October 20120–30Calcic ChernozemNovi Sad, Serbia[106]
1 Experiment only focused on short- and medium-term effects (2-year-long experiment). 2 For AL in the Ap horizon, which was close to 30 cm. For GL in the A horizon, which was on average 28.5 cm. Not available data was denoted as n. a.
Table 2. Summary of results measuring the effect of land management on nitrogen (N) dynamics, in particular total N, available N, N leaching, and nitrous oxide (N2) emissions, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
Table 2. Summary of results measuring the effect of land management on nitrogen (N) dynamics, in particular total N, available N, N leaching, and nitrous oxide (N2) emissions, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
PropertyRelationNumber of SitesSampling StrategySampling TimeDepth [cm]Soil Type/Texture ClassStudy RegionReference
Total NGL = AL2paired sites, repeated samplingJune 2006–June 20070–100Cumulic PhaeozemOttawa County, Kansas, USA[82]
8paired sites, chronosequenceMarch and July 2010 and 20110–10Silty loam, Brown EarthSoutheastern Scotland[62]
GL > AL2paired sitesJuly 2004–May 20050–30 n. a.Northeast Pakistan[73]
1repeated sampling1991–20120–15Dystric CambisolAberdeen, NE Scotland[85]
2paired sites, repeated samplingOctober 20080–75Chromic Luvisol (silty clay loam)Rothamstedt, Harpenden, UK[53]
12paired sitesn. a. 0–30Gleyic/Haplic Leptosols, Haplic/Stagnic CambisolsOuter Western Carpathians, Poland[83]
22chronosequence, repeated sampling1945–1970n. a.VaryingEast central England[132]
2paired sitesNovember 1984–May 19860–28ClaySouth-East Queensland, Australia[164]
313chronosequence1995–20010–10n. a.New Zealand[98]
160paired sites, chronosequenceMarch and June 20150–15Lithology: Karst dolomite and limestone, non-karst clasoliteSouthwest China[166]
717chronosequence2000 and 2004top horizonCambisols (mostly), Leptosols, Regosols, Stagnosols, Albeluvisols, Planosols, GleysolsBavaria, Germany[101] 1
6chronosequenceJuly 20120–10Calcic-Orthic Aridisols (sandy loam)Inner Mongolia, Northern China[102]
Available NGL = AL8paired sites, chronosequenceMarch and July 2010 and 20110–10Silty loam, Brown EarthSoutheastern Scotland[62]
n. a.review (chronosequence, repeated sampling) n. a.n. a.n. a.Worldwide[63]
398 (81 studies)meta-analysisn. a.0–100+n. a.Worldwide[81]
N leachingGL < AL4paired sites, repeated samplingApril 2014–May 20150–10Histic Gleysol, Plaggic AnthrosolLower Saxony, Germany[36]
n. a.reviewn. a.n. a.Varyingn. a.[60]
2repeated samplingOctober 2010–September 20150–15Silt, loamCork, Ireland; Brittany, France[167]
250meta-analysis1991–2009n. a.Sandy soilsNetherlands[168]
6paired sitesApril 2005–June 2012105CambisolLusignan, France[35]
31paired sites, repeated sampling1986–2005n. a.Mostly Cambisol, Planosol, GleysolCesky Krumlov district, Czech Republic[169]
62paired sites, repeated sampling2003–20060–30Eutric FluvisolJena, Germany[170]
82paired sites, repeated sampling2003–20070–15Eutric FluvisolJena, Germany[171]
2paired sitesNovember 1984–May 19860–28ClaySouth-East Queensland, Australia[164]
N2O emissionsGL < AL2paired sites, repeated samplingJune 2002–September 20040–20AlfisolSouthwest Michigan, USA[172]
6paired sites2007–20080–10Silty clay loamQuébec City, Canada[52]
4paired sites2010–20110–10Eutric LuvisolKiel, Germany[173]
GL = AL6paired sites, repeated samplingMay–July 20140–30Histic Gleysol, Plaggic AnthrosolLower Saxony, Germany[174]
GL > AL6paired sitesApril 1997–February 19980–15Loamy sand, sand, silty loam, sandy loamBelgium[33]
1 N stocks were similar, yet N concentrations were higher under grassland compared with under arable land. Not available data was denoted as n. a.
Table 3. Summary of results measuring effects of land management on soil phosphorus (P) content, in particular available and total P, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
Table 3. Summary of results measuring effects of land management on soil phosphorus (P) content, in particular available and total P, with the comparison of a given property simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
PropertyRelationNumber of SitesSampling StrategySampling TimeDepth [cm]Soil Type/Texture ClassStudy RegionReference
Available PGL < AL6chronosequence2008–20090–30Silt, clay, clay loam, silty loamWellesbourne, England[15]
GL = AL8paired sites, chronosequenceMarch and July 2010 and 20110–10Silty loam, Brown EarthSoutheastern Scotland[62]
GL > AL2paired sitesn. a. 0–15 n. a.Northeast Pakistan[72]
2paired sitesJuly 2004–May 20050–30 n. a.Northeast Pakistan[73]
Total PGL < AL6chronosequence2008–20090–30Silt, clay, clay loam, silty loamWellesbourne, England[15]
160paired sites, chronosequenceMarch and June 20150–15Lithology: Karst dolomite and limestone, non-karst clasoliteSouthwest China[166]
GL = AL2paired sites, repeated samplingJune 2006–June 20070–100Cumulic PhaeozemOttawa County, Kansas, USA[82]
64paired sites, chronosequenceMarch and April 20070–20VaryingEngland[205]
8paired sites, chronosequenceMarch and July 2010 and 20110–10Silty loam, Brown EarthSoutheastern Scotland[62]
Not available data was denoted as n. a.
Table 4. Summary of results measuring effects of land management on soil pH levels, with the comparison simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
Table 4. Summary of results measuring effects of land management on soil pH levels, with the comparison simplified into categories as follows: being larger on grassland (GL > AL), larger on arable land (GL < AL) or not significantly different (GL = AL).
PropertyRelationNumber of SitesSampling StrategySampling TimeDepth [cm]Soil Type/Texture ClassStudy RegionReference
pH levelGL < AL2paired sitesn. a. 0–15 n. a.Northeast Pakistan[72]
2paired sitesJuly 2004–May 20050–30 n. a.Northeast Pakistan[73]
1repeated samplingn. a.0–20Clay loamIndagai Mountain Pass, Cankiri, Turkey[86]
4paired sitesAugust 1999–June 20010–10Typic Hapludoll (coarse-loamy)Córdoba, Argentina[215]
3009meta-analysis2011–20180–50VaryingGermany[213]
6paired sitesn. a. 0–30Vertic CambisolCalabria, Italy[31]
313chronosequence1995–20010–10n. a.New Zealand[98]
4paired sitesOctober 2002–October 20040–10Sandy loamMelle, Belgium[100]
717chronosequence2000 and 2004top horizonCambisols (mostly), Leptosols, Regosols, Stagnosols, Albeluvisols, Planosols, GleysolsBavaria, Germany[70]
GL = AL4paired sitesMay 20040–50Stagnic Vertisol, ArenosolThuringia, Germany[78]
6chronosequence2008–20090–30Silt, clay, clay loam, silty loamWellesbourne, England[15]
Not available data was denoted as n. a.
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Filipiak, M.; Kuka, K. Comparison of Chemical Soil Properties of Temperate Grassland and Arable Land—A Review. Soil Syst. 2026, 10, 20. https://doi.org/10.3390/soilsystems10010020

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Filipiak M, Kuka K. Comparison of Chemical Soil Properties of Temperate Grassland and Arable Land—A Review. Soil Systems. 2026; 10(1):20. https://doi.org/10.3390/soilsystems10010020

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Filipiak, Matthias, and Katrin Kuka. 2026. "Comparison of Chemical Soil Properties of Temperate Grassland and Arable Land—A Review" Soil Systems 10, no. 1: 20. https://doi.org/10.3390/soilsystems10010020

APA Style

Filipiak, M., & Kuka, K. (2026). Comparison of Chemical Soil Properties of Temperate Grassland and Arable Land—A Review. Soil Systems, 10(1), 20. https://doi.org/10.3390/soilsystems10010020

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