Open access peer-reviewed chapter

Soil Degradation Processes Linked to Long-Term Forest-Type Damage

Written By

Pavel Samec, Aleš Kučera and Gabriela Tomášová

Submitted: 25 April 2022 Reviewed: 07 July 2022 Published: 02 August 2022

DOI: 10.5772/intechopen.106390

From the Edited Volume

Forest Degradation Under Global Change

Edited by Pavel Samec

Chapter metrics overview

165 Chapter Downloads

View Full Metrics


Forest degradation impairs ability of the whole landscape adaptation to environmental change. The impacts of forest degradation on landscape are caused by a self-organization decline. At the present time, the self-organization decline was largely due to nitrogen deposition and deforestation which exacerbated impacts of climate change. Nevertheless, forest degradation processes are either reversible or irreversible. Irreversible forest degradation begins with soil damage. In this paper, we present processes of forest soil degradation in relation to vulnerability of regulation adaptability on global environmental change. The regulatory forest capabilities were indicated through soil organic matter sequestration dynamics. We devided the degradation processes into quantitative and qualitative damages of physical or chemical soil properties. Quantitative soil degradation includes irreversible loss of an earth’s body after claim, erosion or desertification, while qualitative degradation consists of predominantly reversible consequences after soil disintegration, leaching, acidification, salinization and intoxication. As a result of deforestation, the forest soil vulnerability is spreading through quantitative degradation replacing hitherto predominantly qualitative changes under continuous vegetation cover. Increasing needs to natural resources using and accompanying waste pollution destroy soil self-organization through biodiversity loss, simplification in functional links among living forms and substance losses from ecosystem. We concluded that subsequent irreversible changes in ecosystem self-organization cause a change of biome potential natural vegetation and the land usability decrease.


  • global environmental change
  • pollution
  • nitrogen deposition
  • deforestation
  • soil self-organization

1. Introduction

Human activity induces degradation of many ecosystems, of which the forest degradation results in far-reaching alterations in nutrient cycles in other related types of the environment. The forest degradation, including impacts on ecosystem functions, is intensified by terrestrial environmental change. Forests are most affected by felling and critical loads of pollution. Both processes may be characterized by negative impacts on soil which subsequently causes a decline in the forest natural ability to regenerate [1]. The damage of forest soils signifies that the development of potential natural vegetation is endangered. Damaged forest soils do not allow restoration of original plant community due to disturbed mitigation of environmental fluctuations. Soil capability to mitigate environmental fluctuations resides in uninterrupted cycles of organic matter and continuous fertility. The disruption of forest soil nutrient cycles disadvantages management utilization including sustainable landscape management [2].

Global environmental change involves a related sequence of biophysical, ecosystem and socio-economic alterations that damage life-sustaining abilities of the planet [3]. The current global change is caused by human transformation of the natural environment, but also by the reaction of human communities to the induced modifications. Human transformations of natural environment are concentrated in a critical zone. The critical zone is range among sphere interfaces on the Earth’s surface, where human changes in structure, chemical composition, radiation balance and biodiversity extend [4]. The most vulnerable part of the terrestrial critical zone is composed of soil (pedosphere), which includes interfaces among atmosphere, hydrosphere, lithosphere and biosphere. For these reasons, the soil damage has been altering character of the entire ecosystem over a long period [5].

The most serious soil damage is due to land use modifications after deforestation. While mere forest felling only gives rise to reversible changes in ecosystem functions, the combination of forest felling with soil erosion or conversion for the need of subsequent management utilization generates irreversible ecosystem degradation. Deforestation is instantly followed by declining evaporation and soil loss. The imminent consequences of deforestation are gradually leading to the regional climate change, the loss of ecosystem recoverability and uninhabitable landscape [6]. Regional climate change is mainly caused by reduction of water cycle between the lower-lying areas with higher evaporation and the higher-lying areas with higher atmospheric precipitation in the catchment. The evaporation reduction after deforestation is not sufficient to create cloudiness to make surfaces cooler. Subsequent decrease in precipitation over the higher-lying parts of the river basin deepens water shortage as well as further evaporation decrease in the lower-lying parts [7]. The forest ecosystem recoverability loss is caused mainly due to depletion of the organic matter from the exposed soils, which stimulates germination of tree seeds by means of hormonal effects and moisture retention [8]. Ultimate landscape uninhabitability is caused by uncontrollable soil erosion as a result of surface exposure to wind and landslides of weathered rocks impoverished of organic binders [9].

Forest ecosystem restoration is made impossible within the recent global change, except for soil erosion by pollution. Nitrogen pollution from industry and agriculture has become a major environmental driver of the forest growth [10]. However, atmospheric pollution with the available nitrogen forms is manifested contradictorily within different soil types in forests. The forests situated on optimally fertile soils were generally favorably affected by nitrogen pollution while the predominant forests located on poor soils were damaged. On the one hand, adequate nitrogen intake supports plant growth and, on the other hand, it increases demands on other mineral resources which are declining as a result of human changes in the environment [11]. The unnaturally increasing disparity between plant demands and dwindling nutrient resources causes growth decrease and gradual ecosystem degradation even in hitherto unspoilt areas [12]. Even though the largest nitrogen deposition occurs in the vicinity of pollution sources with lower precipitation, higher concentrations of available nitrogen in wet deposition acidify ecosystems significantly. Approximately 70–80% of nitrogen released from industrial products falls back to the Earth’s surface [13]. Of the nitrogen inputs, 5% penetrates the groundwater, 12% is released into the atmosphere, 30% is immobilized in soil organic matter and 53% is removed with the crop. The utilization of nitrogen by the plant production is still declining, whereas the rate of nitrogen losses by leaching and gasification as well as immobilization in the soil increases in proportion to the amount of fertilizers [14]. Nitrogen supplied to the soil by means of fertilizers results in faster depletion of available bases, making the soil more susceptible to acidification [15].

The environmental nitrogen load is becoming an increasingly important driver of the global ecosystem change as it has exceeded the critical level in large areas of most continents [16]. Exceeding critical nitrogen loads extended plant susceptibility to drought [2, 17, 18]. The widespread plant susceptibility is compound of growing sensitivity of terrestrial ecosystems to climate change. Subsequently, the processes of the climate change and alterations in complex growth conditions for plant communities lead to a deviation in development of prospective natural vegetation or to biome alteration [19]. Therefore, the soil protection is becoming a tool to mitigate the effects of the global terrestrial change maintaining ecosystem link among forests, water cycle and human civilization [5].


2. Forest soil degradation processes

2.1 Impacts on self-organization

Soil degradation may be ranked to one of the most dangerous human activities on the Earth’s surface because soil is not instantly renewable. Degradation commences with vegetation coverage damage as a result of which evaporation decreases. The evaporation diminution foreshadows regional warming which contradictorily results in an intensification of water cycle into short, intense rainfall more frequently accompanied by soil drift or even flash floods [20]. Soil degradation affects ecosystem self-organization. The disruption of soil self-organization is initiated with the decrease in the diversity of functional connections within microbial communities. Disintegration of soil functional interconnectedness involves biodiversity loss and substitution of symbioses for decomposers (saprophytes) that do not exchange available nutrients among organisms, but cause leakage of substances from the ecosystem [21]. Forest soil degradation destroys irreplaceable natural values that improve adaptability of cultural landscape to climate change [22]. The disruption of soil self-organization damages both the continuity of crop production and success of ecosystem restoration.

Soil degradation is divided into quantitative or qualitative one (Figure 1). Quantitative soil degradation represents the physical loss of a soil body. Qualitative soil degradation involves unfavorable alterations in soil physical or chemical properties that limit ecosystem functions [23]. Soil losses occur through claim, erosion or desertification:

  • Claim is usually accompanied by soil sealing, where the land is either merely covered or removed and replaced with building materials. The claim completely destroys soil infiltration capacity. As a result of clearing, the radiation balance and heat capacity alter locally and surface runoff increases sharply [24].

  • Soil erosion is surface soil drift by gravitational shifts, water or wind. The human activity has intensified soil erosion after vegetation removal, excessive grazing and inappropriate tillage, developments affecting landscape and pollution which stopped formation of organo-mineral particles aggregating the soil into more cohesive peds (Figure 2). The vulnerability through erosion (erodibility) depends on weatherability of soil-forming substrate, soil cohesion, climate and land use (Table 1).

  • Desertification is unnatural spread of wastelands after permanent vegetation removal. Causes of unnatural desertification are mainly disproportionate grazing, fires and erosion followed by loss of soil water retention capacity. Deserts spread the fastest in areas naturally adapted to seasonal drought [25]. Approximately 10–20% of the world’s semi-deserts and steppes are threatened by desertification. The accompanying phenomena of desertification are decrease in groundwater levels or salinization which make it impossible to restore vegetation and lead to wasteland homeostasis [9].

Figure 1.

Division of soil degradation processes along quantitative and qualitative impacts on physical or chemical properties.

Figure 2.

Coupled occurences of soil erosion and spreading of desert in arid environments seriously threat forest restoration due to water availability decrease.

ErodibilityParent rockSoil units
Very easyEolian depositsLeptosols, Retisols, Anthroposols
EasyClayey shales, basaltic tuffsLeptosols, Luvisols, Cambisols
MediumCarbonate deposits, sandstonesCalcaric Leptosol, Chromic Cambisol, Ferralic Podzol
Medium hardBreccias, graywackes, phyllites, andesitesStagnic Podzol, Stagnosols, Stagnic Cambisol
HardMetamorphites, basalts, dioritesGleysols, Stagnic Luvisol, Haplic Cambisol
Very hardIgneous rocks, shalesEpi-humic Cambisol, Vertic Chernozem, Histosols

Table 1.

The vulnerability of forest soils by erosion (erodibility) along various parent rocks and developed soil units.

Qualitative soil degradation is produced by excessive losses of the organic matter, the reduction of the biological activity, acidification, contamination, technological compaction (pedocompaction), technical or wind salinization and technical modifications of soil properties. Although qualitative soil degradation potentially occurs in much smaller areas than quantitative one and its effects are usually reversible, they have a similar overlap on landscape functions in the form of reduced water retention capacity and biodiversity, increased runoff and substance imbalances.

Degradation of forest soils is distinctive mainly to the qualitative damage. Qualitative degradation in forests prevails due to the long-term growth of continuous tree species communities. Tree species instantly impede quantitative damage to soils; on the other hand, periodic windthrows during storms mingle the mass among soil horizons, as well as move the soil down the slope. These post-disturbance movements of earth bodies divide microrelief and homogenize soil properties, but at the same time their arrangement is concentrated along the effects of individual global climate changes [26]. However, predominant qualitative degradation of forest soils is manifested by deterioration of physical or chemical properties of solid soil bodies due to external human activities (Figure 3) [27].

Figure 3.

A series of differently degraded soil bodies in mountain conditions: introskeletal erosion in Dystric Hyperskeletic Leptosol (A); surface scarification of Entic Podzol (B) and accumulated Spolic Garbic Urbic Technosol (C).

Deterioration of forest renewal as a consequence of qualitative soil degradation commences with fertility change. Forest felling is accompanied by accelerated leaching of nitrogen substances which can only be ceased by sufficient available calcium [28]. Leaching is preceded by increase in C/N which indicates decrease in ability of soil organic matter to bind mineral nutrients. Soils damaged by compaction of the profile middle part, texturally significantly differentiated or hydromorphic, are exposed to a slow-motion water flow, which expels air for the root growth and similarly increases C/N [29]. The root systems grow merely shallowly with water stagnation and the nutrient loss, making forest stands more susceptible to soil moisture fluctuations [30]. Thus, felling of forests threatened by qualitative erosion impairs the ecosystem ability to restore as a result of exposure to episodic drought [2].

2.2 Physical degradation

Degradation of soil physical properties includes structural damage, pore loss and compaction. The processes of physical soil damage result in both loss of water retention capacity and humus loss by introskeletal erosion. The decline in soil water retention capacity is usually caused by repeated heavy machinery moving. Heavy machinery moving worsens soil aeration and water permeability. Reduced aeration is reflected in the decrease of blank spaces for plant roots and the consequent reduction in the biological activity.

Forest soils are less endangered by physical degradation than agricultural ones owing to dampening effect of surface humus. Nevertheless, topsoil compaction can initiate introskeletal erosion. The mechanical degradation threat is descending from Histosols and gleyed soils to granularly light drying soil bodies [31]. The risks by mechanical damage to the forest soils vary along the grain-size composition, relief exposure and groundwater level (Table 2) [32].

  • A very high risk results from high groundwater level and incoherent soil-forming substrate. Forests endangered by a very high risk of soil erosion are mostly found on Histosols or Gleysols, but also on water-affected Luvisols, Arenosols or Podzols.

  • A high risk emerges from extremely developed hydromorphic features related to heavy skeletality of soil bodies. Forests exposed to high erosion risk are covering Planosols, Stagnosols and Gleysols, including gleyed subtypes on clayey shales or claystones the most.

  • A medium risk is associated with the medium level of forest site gleyfication. Forests at the medium soil erosion risk are located on Stagnic Cambisols, Stagnic Luvisols or Stagnic Fluvisols.

  • A moderate risk is related to site desiccation. Forests moderately endangered by erosion means may be found on Cambisols, Luvisols, Chernosols or on Fluvisols developed from sandy substrates.

  • An insignificant risk is conditioned by soil cohesion, medium skeletability and merely slightly by sloping relief. Forests insignificantly exposed to soil erosion occur at unexposed sites constituted by Leptosols, Cambisols, Podzols or by Chernosols.

RiskSoil seriesInprint depthConsistencyCritical pressure (kPa)
Dry conditionsWet conditions
Very highHistosols+Gleysols≥35Cohesionless30–505–12
HighStagnosols+stagnic soil groups25–35Viscous50–14012–22
MediumStagnic Cambisols-Stagnic Luvisols-Fluvisols15–25Crombly140–30018–50
Moderateunhydromorphic Cambisols+Luvisols+Chernozems+Regosols7–15Cohesive300–60050–80

Table 2.

Characteristics of forest soil compaction risk after logging machinery movement.

Introskeletal erosion represents a predominantly vertical subsidence of fine-grained soil particles through blank spaces among skeleton to the base of rock mantle. The introskeletal erosion risk resides in unstable occurrence of surface humus. Introskeletal erosion is triggered after removal of the vegetation cover in the exposed sites. Its result is the loss of whole fine-grained matter, followed by impossibility of restoring plant community and permanent exposure of relief [33]. The threat to the site by introskeletal erosion is distributed along exposure of relief and soil skeletability [34]:

  • An extreme risk accompanies periglacial brash in arcto-alpine conditions. Extremely endangered sites are only merely sparsely populated by forests. Emerging plant communities are very sensitive to any changes in growth conditions, so they require consistent protection.

  • A very strong risk accompanies shallow soils. Forests endangered by very strong risk of introskeletal erosion are most frequently found along upper tree vegetation limit that is sensitive to global warming [35].

  • A strong risk is accompanied with brash or stone fields (Figure 4). Forests exposed to the strong risk of introskeletal erosion are mostly concentrated on long rocky slopes below the upper limit of tree species vegetation.

  • A medium risk is characteristic of islet occurrences of brash on rocky slopes. Forests exposed to the medium risk of introskeletal erosion typically occur in the middle parts of mountain ranges.

  • A low risk is specific for sparse outcrops of subsoil decay on medium rocky slopes. Forests threatened by erosion on a small scale occur on gentle slopes with deeply developed soils

Figure 4.

Introskeletal erosion leaves rocky flows without surface humus instead soil, where plants can to root hardly, thus forest site gets features of disperse platforms with dwarf vegetation.

2.3 Chemical degradation

2.3.1 Acidification

Degradation of soil chemical properties is the intensification of naturally processed weathering and substance leaching. Chemical soil degradation includes acidification, salinization and intoxication. Acidification is the most extensive process of forest soil degradation causing decline in site fertility [36]. Soil acidification is gradual decreased in neutralizing capacity. In nature, acidification is elicited mainly by water autoprotolysis, naturally acid atmospheric precipitation, organic acids activities, but also by formation of strong acids after reactions of water with atmospheric gases (CO2, SO2) or with some rock-forming minerals (chlorides, sulphates or carbonates). The resulting acids (formal HCl, H2CO3 and H2SO3) can cause very intensive decomposition of original minerals into salts [37].

The intensification of soil acidification was caused by fertilization, crop cultivation and industrial pollution. Industrially emitted CO2, SO2 and NOx create formal acids and soil bases are excessively depleted to neutralize them. The base loss slows down humus formation; on the other hand, raw humus is a significant source of organic acids. The slow-motion formation of humus is reflected in decrease of organo-mineral colloid genesis as a result of which the number of binding sites for exchange cations on the active surfaces of soil particles decreases. The decomposition of variable organo-mineral colloids limits base cations exchanges to stable mineral colloids. Nonetheless, mineral colloids can capture only 0.2–25% of exchangeable cations, unlike organic particles [38].

The soil resists acidification impacts by exchange reactions between inputs of acid-forming H3O+ and available sources of releasable cations. Soil cation sources are active depending on pH (Figure 5). Intensified acidification of forest soils is naturally slowed down either by deciduous tree species or by weathering of the soil-forming substrates. The influence of tree species predominates in surface soil horizons while the influences of soil-forming substrate predominate in the subsurface horizons. Significant acidification in surface horizons of forest soils most often affects the transitional ecosystem types [39]. Even though the mitigating effect of tree species does not overcome impacts of weathering, the optimal tree species composition actively reducing C/N prolongs weathering effects. On the other hand, soil cation release by weathering maintains intensified acidification as a reversible process. Weathering counteracts acidification by means of electrochemically controlled soil-forming substrate decomposition [12, 40, 41]:

  1. The carbonate zone (pH > 6.8) is employed by dissolving the CO32− compounds, whereby the incoming H+ is neutralized into soluble salts. The consequence of these acid–base reactions is a gradual loss of carbonates by dissolution and leaching, which can only be forestalled in soils formed directly from carbonate substrates and rocks.

  2. The silicate zone (pH 5.0–6.8) occurs either within soils from which carbonates have already been leached or where silicates predominate. Acids cause decomposition of silicates from which base cations are released and deuterogenous clay minerals are formed.

  3. The exchange zone (pH 4.2–5.0) may be found in those soils where there is a disproportion between base cations released during weathering of silicates and H+ inputs. The excess of entering H+ is trapped on surface of organic colloids to release bases.

  4. The aluminum zone (pH 3.0–4.2) subdues effects of acidic inputs by releasing Al3+ in the presence of sesquioxides with simultaneous formation of organic complexes. Soil fertility decreases, nutrients are further leached and the biological activity decreases.

  5. The iron zone (pH < 3.0) occurs in those soils where acidic inputs are subdued by the dissolution of iron oxides, Fe3+ migration and destruction of clayey minerals. Nutrients are excessively leached out from these soil bodies, the concentration of toxic substances in soil profile increases and the biological activity is usually concentrated only into raw surface humus.

Figure 5.

Intervals of soil acidity (pH) and organic matter C/N ratio divide trophic (A – Oligotrophic; AB – Oligomesotrophic; B – Mesotrophic; BC – Mesotrophically nitrophilous; C – Nitrophilous; CD – Nitrophilous-base; BD – Mesotrophically base; D – Base) series among zones buffering acidification through specific neutralization. Data according to [39].

The unnatural decomposition of soil minerals triggers irreversible acidification. Acidification may be mitigated merely after the removal of acidifying substances sources. The acidification of forest soils affected exchange zone the most, switching to active aluminum zone [42]. The damage to the forest ecosystem by release of active Al3+ followed due to occupation of exchange sites on soil particle surfaces instead of bases, the lack of which limited root growth. The roots were concentrated shallowly below the surface so that new focal points of biological activity and humus ceased to form deeper in the soil [43]. Introducing the other side of the fact, the marginally widespread transition from the aluminum zone to iron one was ensued by loss at the ability of forest ecosystems to restore from the damage (Figure 6).

Figure 6.

Irreversible damaged forests are characteristic by predominantingly dead tree storey and by absent young woods due to lost soil organic matter irreplaceably stabilizing moisture during seed germination.

Air pollution has significantly accelerated soil acidification, especially in the areas of forests transformed into homogeneous stands of coniferous tree species. While cultivation of homogeneous coniferous forests homogenized formation of acidic humus causing micropodzolization and increased base cation leaching, the pollution after acid deposition reduced not only the forest increment but also decomposition of organic matter [44]. Forest increment was reduced by direct damage to the assimilation apparatus, by stimulating sensitivity to seasonal drought or frost and by reduction in soil symbioses mediating nutrient deficiencies. The decline of mycorrhizal fungi was followed by increase in frequency of saprophytic to saproparasitic fungi, which diverted organic matter decomposition to complete leaching from the ecosystem [45]. The susceptibility of mycorrhizal symbioses to pollution resulted in limited accessibility to phosphorus necessary for nucleic acid synthesis [46]. The disturbed phosphorus cycle triggered decrease in increment as well as seed germination leading to forest self-organization loss [47].

2.3.2 Salinization

Soil salinization is the process of accumulating surpluses of mineral salts. Salinization of forest soils is a rare phenomenon, but it threatens 23% of agricultural land, mainly in arid areas [48]. Forest soils are salinized in areal or linear extent. Areal salinization is caused by high groundwater mineral levels, the use of saline water for irrigation, waste materials for fertilization or deposition of solids. Linear salinization occurs alongside roadsides maintained by chemical salting during winter or along river banks. The recent climate change is expanding areas of salinized soils with rising sea level along the coast or estuaries. On the other hand, the natural risk of soil acidification subdues consequences of salinization [49].

The impacts of salinization in forests are associated with extreme soil chemical properties. Salinization highlights malfunctions of water and nutrient uptake by plants. Above all, the disproportionate sodium input (sodification) disrupts ration among exchangeable bases in the soil environment, thereby disrupting effects of alkalization on soil structure. Significant Na+ inputs displace other cations from soil sorption complex and disperse soil particles. Sodium displacement of cations results in deficient nutrition, but at the same time crushes soil structure, thus water availability fluctuates. Sodium surplus in plant tissues reduces osmotic pressure, whereby cells lose ability to absorb other substances from soil solution [27]. While conifers are susceptible to soil salinization, deciduous tree species are tolerant to it. The younger plants are more susceptible than the older ones.

Areal forest salinization is most at risk in floodplains due to variability of water flow. The regulation of water flow caused groundwater level fall in some river basins while it resulted to water level permanent increase in some other ones. The groundwater level decline was typically ensued by ecosystem desiccation due to the fact that riparian forests are mostly located in submontane locations with insufficient precipitation [50]. By contrast, rising groundwater levels after water regulation meant change in availability of mineral ions, with impacts on soil microbial activity and ability of the ecosystem to sequester carbon. The increase in level of saline water inflicts decrease in soil microbial activity and consequently decrease in vegetation growth [51]. On the one hand, decreases in growth processes are caused by loss of oxygen in soil environment, on the other hand, by increasing concentrations of Na+, Cl, SO42− ions, including Fe and Mn compounds. In particular, SO42− in the soil solution is converted to toxic sulphide when there is a lack of oxygen. Although Fe and Mn are biogenic elements that catalyze soil organic matter decomposition, bound in sulphides block microbial metabolism [52]. The main forest salinization danger with groundwater is inability to adapt on climate change due to spread of microaerobic conditions [53, 54].

2.3.3 Intoxication

Intoxication of soils with heavy metals, radioactive or petroleum substances is a rare but very hazardous process of cumulative pollution. Especially, heavy metals merely slowly participate in biogeochemical cycles and accumulate in the ecosystem because they are either microbiogenic (Cu and Zn), or xenobiotic (e.g. Cd, Co, Pb, Hg, Ni).

Soil intoxication occurs by deposition means. Sources of cumulative pollution are point or dispersed ones. The point sources of heavy metals are smelters, thermal power stations or municipalities by watercourses. The dispersed sources are represented by polluted water, inappropriate distribution of industrial or sewage sludge or operation of internal combustion engines. Fluvisols, which are among the most intoxicated soils, are usually located between watercourses and agricultural soils [48].

Xenobiotics are toxic to most organisms. They mainly affect energy balance of living cells and their division. Heavy metals mainly bring about halting respiration as a consequence of interactions with SH- groups at intracellular enzymes and their complexes, they disrupt semipermeability of cell membranes and their proton gradient. Soil environment pollution with heavy metals significantly reduces density of microbial occurrences and directly damages plants. The rate of soil contamination damages microbial activity significantly more than the differences in heavy metal contents among different sites [55]. However, the decline of susceptible species is being replaced by expansion of resistant species populations, including pests abundantly infesting damaged plants [56].

Resistant phytophagous arthropods adapt to the environment contaminated with heavy metals by searching for less contaminated nourishment, sufficient release of metals in excrements, by sloughing or by means of other tissues (in the adipose body, epithelium of the digestive tract) [57]. The importance of phytophagous insects for the movement of heavy metals in ecosystem lies in the fact that these invertebrates are an important link in food web that receive toxic substances directly from plants, especially Cd and Zn [58, 59], but no Ni and Fe [60].

The risk of heavy metal accumulation with toxic manifestations threatens not only tree species, but also predators. Numerous ground beetles (Carabidae) and ants (Formicidae) are mainly food-bound to phytophagous insects, earthworms, countless larvae and springtails [61]. On the other hand, the immobilization of heavy metals takes place in cells of specialized metallogeneic microorganisms by binding to metallothionein-based amino acids. Immobilization of heavy metals in amino acids changes course of humus formation. The rate of biosorption on the soil microbial active surfaces typically decreases in order Zn > Cd > Pb > Cu > Cr [62]. Alterations in the forest ecosystems as a result of heavy metal pollution include (Figure 7):

  • the reduction of phospholipid fatty acids content in the bacterial cells and raw humus. Even though the total content of phospholipid acids is directly proportional to ATP synthesis, it decreases under the toxic load. The consequence of these processes is alterations in overall functional diversity of microbial community and inability to decompose deposition by aromatic hydrocarbons [63].

  • promoting release of mobile humus substances and decrease of insoluble substances. On the other hand, the great ability of Pb to form complexes with insoluble humus substances maintains its immobilization while Cd and Zn tend to form soluble complexes. Although more heavy metals are bound to soluble humus substances than to insoluble organic compounds, the greater release of soluble compounds also contributes to pollutant migration in the soil and to their penetration into groundwater [64].

Figure 7.

Forests dying at regions loaded by acid deposition were transformed to substitute stands of resistant introduced tree species which have provided cover for regeneration of indigenous forest communities after pollution decrease.

Petroleum products belong to secondary persistent organic compounds, similar to benzo(a)pyrene, polychlorinated dibenzo-p-dioxins or dibenzofurans, which are removed from the soil for more than 2 years [65]. Like benzo(a)pyrene, they consist of polycyclic aromatic hydrocarbons that directly harm the health of organisms. Oil pollution is significantly more caused by human activity than by natural (geogenic) sources. It begins with mining, combustion of petroleum products (fossil fuels), accidental or operational spills and corrosion of industrial materials. Oil products in the forest ecosystem load surface humus the most, which at the same time prevents their penetration into deeper occurring soil. The load of surface humus decreases activity of soil microorganisms. The decrease of soil biological activity is mostly caused by aromatic nuclei imitating lignin, which either block formation of amino acids or replace carbon compounds in fungi [66]. Subsequently, the humus decomposition is disrupted, foreshadowing disruption of processes to get available nutrients from the soil. Nevertheless, the load of petroleum substances is irregularly concentrated in surroundings of industrial areas and vertically along different intensities of wood logging in floodplains, hillycountries, highlands and high-mountain forests [67].



The study was supported by projects LM2018123 CzeCOS of the Ministry of Education, Youth and Sports of the Czech Republic and 952314 ASFORCLIC of the European Union’s Horizon 2020 Programme for Research & Innovation.


  1. 1. Hansen MC, Potapov PV, Moore R, Hancher M, Turubanova SA, Tyukavina A, et al. High-resolution global maps of 21st-century forest cover change. Science. 2013;342:850-853. DOI: 10.1126/science.1244693
  2. 2. Bernal S, Hedin LO, Likens GE, Gerber S, Buso DC. Complex response of the forest nitrogen cycle to climate change. Proceedings of the National Academy of Sciences of the United States of America. 2012;109:3406-3411. DOI: 10.1073/pnas.1121448109
  3. 3. Brovkin V, Brook E, Williams JW, Bathiany S, Lenton TM, Barton M, et al. Past abrupt changes, tipping points and cascading impacts in the earth system. Nature Geoscience. 2021;14:550-558. DOI: 10.1038/s41561-021-00790-5
  4. 4. Richter DB, Mobley ML. Monitoring Earthʼs critical zone. Science. 2009;326:1067-1068. DOI: 10.1126/science.1179117
  5. 5. Kučera A, Samec P, Bajer A, Skene KR, Vichta T, Vranová V, et al. Forest soil water in landscape context. In: Datta R, Meena RS, editors. Soil Moisture Importance. London, UK: InTechOpen; 2021
  6. 6. Woo SW. Forest decline of the world: A linkage with air pollution and global warming. Africal Journal of Biotechnology. 2009;8:7409-7414
  7. 7. Lawton RO, Nair US, Pielke RA Sr, Welch RM. Climatic impact of tropical lowland deforestation on nearby montane cloud forests. Science. 2001;294:584-587. DOI: 10.1126/science.1062459
  8. 8. Lucier A, Ayres M, Karnosky D, Thompson I, Loehle C, Percy K, et al. Forest responses and vulnerabilities to recent climate change. In: Seppälä R, Buck A, Katila P, editors. Adaptation of Forests and People to Climate Change – A Global Assessment Report. Vienna, Helsinki: IUFRO; 2009. pp. 29-52
  9. 9. Margono BA, Potapov PV, Turubanova S, Stolle F, Hansen MC. Primary forest cover loss in Indonesia over 2000-2012. Nature Climate Change. 2014;4:730-735. DOI: 10.22146/ijg.12496
  10. 10. Etzold S, Ferretti M, Reinds GJ, Solberg S, Gessler A, Waldner P, et al. Nitrogen deposition is the most important environmental driver of growth of pure, even-aged and managed European forests. Forest Ecology and Management. 2020;458:117762. DOI: 10.1016/j.foreco.2019.117762
  11. 11. Nosengo N. Fertilized to death. Nature. 2003;425:894-895. DOI: 10.1038/425894a
  12. 12. Schröder W, Nickel S, Jenssen M, Riediger J. Methodology to assess and map the potential development of forest ecosystems exposed to climate change and atmospheric nitrogen deposition: A pilot study in Germany. Science ot the Total Environment. 2015;521:108-122. DOI: 10.1016/j.scitotenv.2015.03.048
  13. 13. Galloway JN, Schlesinger WH, Hiran Levy II, Michaels A, Schnoor JL. Nitrogen fixation: Anthropogenic enhancement – Environmental response. Global Biogeochemical Cycles. 1995;9:235-252. DOI: 10.1029/95GB00158
  14. 14. Goulding KWT, Bailey NJ, Bradbury NJ, Hargreaves P, Howe M, Murphy DV, et al. Nitrogen deposition and its contribution to nitrogen cycling and associated soil processes. The New Phytologist. 1998;139:49-58. DOI: 10.1046/j.1469-8137.1998.00182.x
  15. 15. Oulehle F, Evans CD, Hofmeister J, Krejčí R, Tahovská K, Persson T, et al. Major changes in forest carbon and nitrogen cycling caused by declining sulphur deposition. Global Change Biology. 2011;17:3115-3129. DOI: 10.1111/j.1365-2486.2011.02468.x
  16. 16. Gruber N, Galloway JN. An earth-system perspective of the global nitrogen cycle. Nature. 2008;451:293-296. DOI: 10.1038/nature06592
  17. 17. Westling O, Fagerli H, Hellsten S, Knulst JC, Simpson D. Comparison of modelled and monitored deposition fluxes of Sulphur and nitrogen to ICP – Forest sites in Europe. Biogeosciences Discussions. 2005;2:933-975. DOI: 10.5194/bg-3-337-2006
  18. 18. Vacek S, Vacek Z, Remeš J, Bílek L, Hůnová I, Bulušek D, et al. Sensitivity of unmanaged relict pine forest in the Czech Republic to climate change and air pollution. Trees. 2017;31:1599-1617. DOI: 10.1007/s00468-017-1572-0
  19. 19. Parmesan C, Yohe G. A globally coherent fingerprint of climate change impacts across natural systems. Nature. 2003;421:37-42. DOI: 10.1038/nature01286
  20. 20. Blöschl G, Kiss A, Viglione A, Barriendos M, Böhm O, Brázdil R, et al. Current European flood-rich period exceptional compared with past 500 years. Nature. 2020;583:560-566. DOI: 10.1038/s41586-020-2478-3
  21. 21. Abbasian F, Lockington R, Megharaj M, Naidu R. The biodiversity changes in the microbial population of soils contaminated with crude oil. Current Microbiology. 2016;72:663-670. DOI: 10.1007/s00284-016-1001-4
  22. 22. Bergstrom DM, Wienecke BC, van den Hoff J, Hughes L, Lindenmayer DB, Ainsworth TD, et al. Combating ecosystem collapse from the tropics to the Antarctic. Global Change Biology. 2021;27:1692-1703. DOI: 10.1111/gcb.15539
  23. 23. Lal R. Soil quality and sustainability. In: Lal R, Blum WH, Valentine C, Steward BA, editors. Methods for Assessment of Soil Degradation. Boca Raton: CRC Press; 1998. pp. 547-554
  24. 24. Bronstert A, Niehoff D, Bürger G. Effects of climate and land-use change on storm runoff generation: Present knowledge and modelling capabilities. Hydrological Processes. 2002;16:509-529. DOI: 10.1002/hyp.326
  25. 25. Trnka M, Hayes M, Jurečka F, Bartošová L, Anderson M, Brázdil R, et al. Priority questions in multidisciplinary drought research. Climate Research. 2018;75:241-260. DOI: 10.3354/cr01509
  26. 26. Šilhán K, Pánek T, Dušek R, Havlů D, Brázdil R, Kašičková L, et al. The dating of bedrock landslide reactivations using dendrogeomorphic techniques: The Mazák landslide, outer Western Carpathians (Czech Republic). Catena. 2013;104:1-13. DOI: 10.1016/j.catena.2012.12.010
  27. 27. Vavříček D, Kučera A. Základy lesnického půdoznalství a výživy lesních dřevin. Kostelec nad Černými lesy: Lesnická práce; 2017. p. 364
  28. 28. Green MB, Bailey AS, Bailey SW, Battles JJ, Campbell JL, Driscoll CT, et al. Decreased water flowing from a forest amended with calcium silicate. Proceedings of the National Academy of Sciences of the United States of America. 2013;110:5999-6003. DOI: 10.1073/pnas.1302445110
  29. 29. Likens GE, Bormann FH. Biogeochemistry of a Forested Ecosystem. New York: Springer-Verlag; 1995
  30. 30. Knapp AK, Beier C, Briske DD, Classen AT, Luo Y, Reichstein M, et al. Consequences of more extreme precipitation regimes for terrestrial ecosystems. Bioscience. 2008;58:811-821. DOI: 10.1641/B580908
  31. 31. Vavříček D, Ulrich R, Kučera A. Ochrana půdy v těžebně dopravní činnosti. Mendelova univerzita v Brně. 2014
  32. 32. Simanov V, Macků J, Popelka J. Nový návrh terénní klasifikace a technologické typizace. Lesnictví-Forestry. 1993;39:422-428
  33. 33. Šach F, Černohous V. Introskeletová eroze. In: Borůvka L, editor. Pedologické dny. 2002. pp. 129-132
  34. 34. Souček J, Kriegel H, Nárovec M, Šach F. Obnova lesa na lokalitách ohrožených introskeleletovou erozí. Lesnický průvodce. 2010;2:1-35
  35. 35. Wang Z, Lyu L, Liu W, Liang H, Huang J, Zhang QB. Topographic patterns of forest decline as detected from tree rings and NDVI. Catena. 2021;198:105011. DOI: 10.1016/j.catena.2020.105011
  36. 36. Cosby BJ, Ferrier RC, Jenkins A, Wright RF. Modelling the effects of acid deposition: Refinements, adjustments and inclussion of nitrogne dynamics in the MAGIC model. Hydrology and Earth System Sciences. 2001;5:499-517. DOI: 10.5194/hess-5-499-2001
  37. 37. Bolan NS, Curtin D, Adriano DC. Acidity. In: Hillel D, editor. Encyclopedia of Soils in the Environment. New York: Academic Press; 2005. pp. 11-17
  38. 38. Ulrich B. The history and possible cause of forest decline in Central Europe, with particular attention to the German situation. In: EC, UN/ECE, Forest Soil Condition in Europe. Brussel, Geneva: Large-Scale Soil Survey; 1995
  39. 39. Samec P. Indication of forest soil fertility deviations by differences between trophic series and soil buffering: Geobiocoenological approach. Phytopedon(Bratislava). 2016;15:1-11
  40. 40. Fenn ME, Huntington TG, McLaughlin SB, Eagar C, Gomez A, Cook RB. Status of soil acidification in North America. Journal of Forest Science. 2006;52(Special Issue):3-13
  41. 41. Ross DS, Matschonat G, Skyllberg U. Cation exchange in forest soils: The need for a new perspective. European Journal of Soil Science. 2008;59:1141-1159. DOI: 10.1111/j.1365-2389.2008.01069.x
  42. 42. Paoletti E, Schaub M, Matyssek R, Wisser G, Augustaitis A, Bastrup-Birk AM, et al. Advances of air pollution science: From forest decline to multiple-stress effects on forest ecosystem services. Environmental Pollution. 2010;158:1986-1989. DOI: 10.1016/j.envpol.2009.11.023
  43. 43. Heim A, Luster J, Brunner I, Frey B, Frossard E. Effects of aluminium treatment on Norway spruce roots: Aluminium binding forms, element distribution, and release of organic substances. Plant and Soil. 1999;216:103-116
  44. 44. Klimo E, Materna J, Lochman V, Kulhavý J. Forest soil acidification in the Czech Republic. Journal of Forest Science. 2006;52(Special Issue):14-22
  45. 45. Hagerberg D, Thelin G, Wallander H. The production of ectomycorrhizal mycelium in forests: Relation between forest nutrient status and local mineral sources. Plant and Soil. 2003;252:279-290. DOI: 10.1023/A:1024719607740
  46. 46. Rejšek K. Acid phosphomonoesterase activity of ectomycorrhizal roots in Norway spruce pure stands exposed to pollution. Soil Biology and Biochemistry. 1991;23:667-671. DOI: 10.1016/0038-0717(91)90081-T
  47. 47. Talkner U, Meiwes KJ, Potočić N, Seletković I, Cools N, De Vos B, et al. Phosphorus nutrition of beech (Fagus sylvatica L.) is decreasing in Europe. Annals of Forest Science. 2015;72:919-928. DOI: 10.1007/s13595-015-0459-8
  48. 48. FAO-ITPS. Status of the World’s soil resources (SWSR) – Main report. Rome: Food and agriculture Organization of the United Nations. Intergovernmental Technical Panel on Soils. 2015
  49. 49. Shi WM, Yao J, Yang F. Vegetable cultivation under greenhouse conditions leads to rapid accumulation of nutrients, acidification and salinity of soils and groundwater contamination in south-eastern China. Nutrient Cycling in Agroecosystems. 2009;83:73-84. DOI: 10.1007/s10705-008-9201-3
  50. 50. Klimo E, Hager H, Matić S, Anić I, Kulhavý J. Floodplain forests of the temperate zone of Europe. Kostelec nad Černými lesy: Lesnická práce; 2008
  51. 51. Machado RMA, Serralheiro RP. Soil salinity: Effect on vegetable crop growth. Management practices to prevent and mitigate soil salinization. Horticulturae. 2017;3:30. DOI: 10.3390/horticulturae3020030
  52. 52. Lamers LP, Dolle GE-T, Berg STVD, van Delft SP, Roelofs JG. Differential responses of freshwater wetland soils to sulphate pollution. Biogeochemistry. 2001;55:87-101. DOI: 10.1023/A:1010629319168
  53. 53. Herbert ER, Boon P, Burgin AJ, Neubauer SC, Franklin RB, Ardón M, et al. A global perspective on wetland salinization: Ecological consequences of a growing threat to freshwater wetlands. Ecosphere. 2015;6:206. DOI: 10.1890/ES14-00534.1
  54. 54. Pechanec V, Machar I, Kiliánová H, Vyvlečka P, Seják J, Pokorný J, et al. Ranking the key Forest habitats in ecosystem function provision: Case study from Morava River basin. Forests. 2021;12:138. DOI: 10.3390/f12020138
  55. 55. Friedlová M. The influence of heavy metals on soil biological and chemical properties. Soil & Water Research. 2010;5:21-27. DOI: 10.17221/11/2009-SWR
  56. 56. Führer E. Air pollution and the incidence of forest insect problems. Zeitschrift für Angewandte Entomologie. 1985;99:371-377. DOI: 10.1111/j.1439-0418.1985.tb02000.x
  57. 57. Lindqvist L. Acumulation of cadmium, cooper, and zinc in five species of phytophagous insects (Swe.). Enviromental Entomology. 1992;21:160-163. DOI: 10.1093/ee/21.1.160
  58. 58. Roberts RD, Johnson MS. Dispersal of heavy metals from abandoned workings after their transference through terrestrial food chains. Environmental Pollution. 1978;16:293-310. DOI: 10.1016/0013-9327(78)90080-0
  59. 59. Nuorteva P. The Effect of Pollution in Development of Forest Insect Outbreaks. Petrozavodsk: Barents Euro-Arctic Region; 1997. pp. 248-249
  60. 60. Beyer WN, Pattee OH, Sileo L, Hoffman DJ, Mulhern BM. Metal contamination in wildlife living near two zinc smelters. Environmental Pollution. 1985;38:63-86. DOI: 10.1016/0143-1471(85)90094-7
  61. 61. Kula E, Purchart L. The ground beetles (Coleoptera: Carabidae) of forest altitudinal zones of the eastern part of the Krušné hory Mts. Journal of Forest Science. 2004;50:456-463. DOI: 10.17221/4641-JFS
  62. 62. Muter O, Lubinya I, Millers D, Grigorjeva L, Ventinya E, Rapoport A. Cr (IV) sorption by intact and dehydrated Candida utilis cells in the presence of other metals. Process Biochemistry. 2002;38:123-131. DOI: 10.1016/S0032-9592(02)00065-1
  63. 63. Laczko E, Rudaz A, Aragno M. Diversity of anthropogenically influenced or disturbed soil microbial communities. In: Insam H, Rangger A, editors. Microbial Communities. Functional Versus Structural Approaches. Berlin – Heildelberg: Springer Verlag; 1997. pp. 57-67
  64. 64. Borůvka L, Drábek O. Podíl těžkých kovů ve frakcích huminových látek kontaminované půdy. In: Borůvka L, editor. Pedologické dny 2002 – Degradace půdy. ČZU v Praze; 2002. pp. 55-58
  65. 65. Němeček J, Vácha R, Podlešáková E. Hodnocení kontaminace půd v ČR. Praha: Výzkumný ústav meliorací a ochrany půdy; 2010. p. 148
  66. 66. Vácha R, Čechmánková J, Havelková M, Horváthová V, Skála J. Přestup polycyklických aromatických uhlovodíků z půdy do vybraných rostlin. Chemicke Listy. 2008;11:1003-1010
  67. 67. Borůvka L, Sáňka M, Šrámek V, Vácha R, Čechmánková J, Čupr P, et al. Methods for the Forest soils pollution assessment. Lesnický průvodce. 2015;12:1-63

Written By

Pavel Samec, Aleš Kučera and Gabriela Tomášová

Submitted: 25 April 2022 Reviewed: 07 July 2022 Published: 02 August 2022