Soil ecosystem services in urban areas and methods for their assessment using remote sensing

Cities and nature are interdependent. Many of the needs of urban population are fulfilled by ecosystem services, which are typically grouped into four categories: provisioning, regulating, habitat, and cultural services. These include such benefits as food and water supply, climate regulation, biodiversity support, as well as various cultural and recreational values. Soil, through its functions, is an important element of the environment that provides ecosystem services in urban areas. This publication focuses on a review of the literature on soil ecosystem services, with particular emphasis on the urban context. It also highlights how remote sensing tools and spatial planning approaches can be applied to monitor soil ecosystem services in cities and support sustainable urban management. The increasing level of urbanization poses a major threat to soil and its ability to provide ecosystem services. Urbanization worsens climatic conditions in cities, threatens biodiversity by interrupting ecological corridors, increases surface runoff creating flood risks, and contributes to elevated levels of pollution in both soil and air. Given that this process is inevitable, minimizing environmental damage requires well-designed land management approaches. Integrating ecosystem services into urban planning will not only contribute to preserving nature in the best possible condition, but will also enhance the quality of life for city dwellers.

1 Introduction

Landscape resources and ecological values have the ability to fulfil current and future human needs, supported by self-regulating and resilience mechanisms that sustain ecosystem functions (Solon, 2008). These needs are met through ecosystem services. Basic physiological needs are mainly served by provisioning and regulating services (production of food, clothing materials and energy sources, air and water purification and regulation of greenhouse gas emissions). Safety and security needs are covered by habitat and regulating services such as flood mitigation, pest and disease control and habitat stability, while self-realization needs are satisfied by cultural services (Dominati et al., 2010). Urbanization poses a significant environmental threat, leading to irreversible changes in ecosystems and directly affecting the supply of soil-related ecosystem services (Setälä et al., 2014). Land-use changes driven by urban expansion disrupt natural soil functions, leading to soil degradation, pollution, and biodiversity loss.

Nature and cities were long considered separate. However, contemporary urban planning increasingly recognizes their interdependence. Cities rely on natural systems to maintain public health, economic stability, and climate resilience. Urban planners now recognize the role of cities as stewards of the ecosystems on which they depend, highlighting the importance of integrating natural elements into urban landscapes (Seto et al., 2013). In addition, unsustainable exploitation of ecosystem services can lead to severe socio-economic consequences, such as limited access to drinking water, increased spread of diseases, soil degradation, and the financial burden of rebuilding infrastructure after extreme weather events (Millennium Ecosystem Assessment, 2005; Czarnecki and Lewandowska-Czarnecka, 2007). Therefore, a comprehensive approach to ecosystem-based urban planning is essential to mitigate these risks and ensure long-term sustainability. According to the United Nations’ projections, the global urban population is expected to reach nearly 70% by 2050 (United NationsDepartment of Economic and Social AffairsPopulation Division, 2022). Since urbanization is inevitable, well-designed land management strategies are necessary to minimize its environmental impact. Figure 1 highlights the urgency of integrating ecological considerations into city planning.

Figure 1

Figure 1. Global urbanization trend (1950–2050) based on UN DESA World Urbanization Prospects.

This review aims to integrate dispersed scientific and policy-relevant knowledge on the ecosystem services provided by soils and how urbanization affects them, with a focus on spatial planning, remote sensing tools, and soil management strategies. While the literature on ecosystem services is well-developed, relatively few studies connect soil-related services directly with urbanization processes and decision-making frameworks. The present review aims to fill this gap by synthesizing evidence across disciplines and scales, and outlining both risks and pathways for action.

To guide this review, we addressed the following key research questions:

1. What ecosystem services are provided by soils in urban environment, and why are they critical for sustainable cities?

2. How do processes of urbanization (e.g., soil sealing, land-use change) affect the capacity of soils to deliver these ecosystem services?

3. How can spatial planning strategies help preserve or enhance soil ecosystem services in urban areas?

4. What tools and approaches does remote sensing offer for monitoring and valuing soil ecosystem services in urban management?

2 Methods and search strategy

2.1 Review design and approach

We carried out an integrative narrative review of peer reviewed studies, complemented by a small set of policy and technical documents, to summarize how urbanization influences soil functions and the ecosystem services they underpin in urban and periurban environments. Because relevant evidence is dispersed across soil science, urban ecology, remote sensing and spatial planning, we focused on building a coherent and well documented evidence base rather than attempting a fully exhaustive inventory. Three themes structured the synthesis: how urbanization pressures (for example, soil sealing, compaction and land take) alter soil properties and functions; how these changes are treated in spatial planning and land management; and how remote sensing contributes indicators for assessment and monitoring. The review questions were framed using PECO logic: the population comprised soils in urban or urbanizing settings; exposure covered urbanization processes; comparators were less urbanized or pre development conditions when reported; and outcomes included soil functions, soil related ecosystem services and their ecological, socio cultural and economic indicators. The evidence base underpinning the synthesis consists of the sources cited in the reference list. The main steps of searching, screening and synthesis are outlined in Supplementary Figure S1.

2.2 Information sources and search strategy

Searches were run in Web of Science Core Collection, Scopus and PubMed/MEDLINE, with Google Scholar used to capture records that are not consistently indexed in the main bibliographic databases. Because evidence relevant to urban soils and planning also appears outside journals, we additionally screened grey literature from institutional repositories and agency websites (including FAO, the European Environment Agency and the European Commission Joint Research Centre), as well as national or regional policy documents when they addressed soil related services in cities. To reduce the risk of missing relevant work, we screened reference lists of key papers and used forward citation searches when they yielded additional relevant records.

Search strings were built around three groups of terms: soil ecosystem services and soil functions; urbanization and urban land use change; and assessment or management terms, including spatial planning and remote sensing. We piloted and refined the strings iteratively, checking that they retrieved benchmark papers and adjusting database syntax to balance sensitivity and specificity. Controlled vocabulary was used where available (for example, MeSH terms in PubMed). A representative query (Web of Science Topic) was: ((“soil ecosystem services” OR “soil function*” OR “soil biodiversity”) AND (urban* OR urbanization OR urbanisation OR “urban development” OR “urban sprawl” OR “soil sealing” OR “land take”) AND (assessment OR mapping OR valuation OR management OR “spatial planning” OR “remote sensing” OR “earth observation” OR GIS)). No date limits were applied. The final search update was completed on 1 August 2025, and retrieved records were exported to a reference manager for deduplication and screening. Full database specific strings are provided in Supplementary Table S1.

2.3 Eligibility criteria

Publications were included in the analysis if they concerned soils studied in a clearly urban or urbanizing context or compared urban soils with suburban or rural soils. They were also included if they discussed, assessed, mapped or evaluated soil functions and ecosystem services in relation to urbanization pressure, planning or management. We considered empirical studies, modelling studies and synthesis papers. Remote sensing studies were included when the remotely derived indicators were interpreted in relation to soil properties or soil mediated processes, rather than being limited to generic land cover mapping. Records were excluded when they focused solely on agricultural production without an urban context, addressed contamination without linking results to soil functions or services, lacked accessible full text, or were not written in English or Polish.

2.4 Study selection and data management

All retrieved records were imported into a reference manager and duplicates were removed. Titles and abstracts were screened against the eligibility criteria, followed by full text assessment of potentially relevant records. Because terminology differs across disciplines, records were retained for full text screening whenever the abstract was ambiguous. Screening decisions were discussed within the author team to keep the interpretation of the criteria consistent, and disagreements were resolved by consensus. At the full text stage, reasons for exclusion were noted to support transparent reporting. An overview of the workflow is provided in Supplementary Figure S1.

2.5 Data extraction and coding framework

Data were extracted using a standardized form that captured bibliographic information and study descriptors (geographic setting, climate context when reported, spatial scale and study design), the main urbanization pressures addressed (for example, soil sealing, land cover change or compaction), the soil properties and indicators used, and the ecosystem services or soil functions discussed. For remote sensing studies we additionally noted the sensor platform, spatial resolution, key preprocessing steps when reported, the indices or products used as proxies and whether validation against field or laboratory measurements was undertaken. The extraction form was piloted on a small subset of studies and refined before full coding. To support consistent interpretation across heterogeneous literatures, we coded the evidence using a cascade logic linking soil properties to soil functions, soil based ecosystem services and, where possible, to societal benefits. Ecosystem services were assigned to the four service classes used throughout the manuscript (provisioning, regulating, habitat and cultural). When services were not named explicitly by the original authors, classification followed the reported outcomes and the described role of soil functions in generating benefits.

2.6 Quality appraisal

Given the range of study designs and disciplines, we applied a pragmatic quality appraisal rather than a single risk of bias tool. For each included record we assessed whether the urban context and soil compartment were described clearly, whether the indicators used for soil functions or services were appropriate and well defined, and whether analytical and spatial methods were reported transparently. For remote sensing studies, particular attention was paid to calibration procedures and to validation against field or laboratory data. This appraisal was used to interpret the strength of the evidence in the synthesis; studies were not excluded solely on quality grounds, but limitations were noted when discussing findings.

2.7 Synthesis and analysis

Synthesis was primarily narrative and followed the structure of research questions. Evidence was grouped by ecosystem service class, by the type of urbanization pressure and by the assessment approach, while keeping track of spatial and temporal scales. Where several studies reported comparable quantitative indicators, results were summarized descriptively, most often as ranges, because definitions and measurement protocols were seldom consistent enough for meta-analysis. Remote sensing approaches were treated as a distinct strand of evidence: we compared data requirements, spatial resolution and the inferential path linking observations to soil related services, and we differentiated indicators that reflect soil properties more directly from those that act as indirect proxies mediated by vegetation or surface characteristics.

2.8 Methodological limitations

This review has limitations typical of multidisciplinary evidence synthesis. Restricting peer reviewed searches to English and Polish may have resulted in missed studies, particularly local case studies that are not indexed. The heterogeneity of definitions and indicators limited the scope for quantitative aggregation, and some topics are supported by small and uneven evidence bases. Remote sensing products, in particular, often represent proxies rather than direct measurements of soil processes; we therefore interpreted them cautiously and highlighted studies that reported validation or triangulation with independent data. Additionally, there may be a tendency in published literature to preferentially include research with definitive or favorable outcomes.

3 Soil functions and ecosystem services

3.1 Soil functions

Soil is a fundamental component of the terrestrial environment. The formation of soil is a long and intricate process, making soil resources effectively non-renewable. Each soil type possesses distinct physical, chemical, and biological properties, which determine its capacity to support plant and animal life (Zawadzki, 1999; Woch, 2015). Soil performs essential ecological and socio-economic functions, which arise from the dynamic interactions between biotic and abiotic components (Morel et al., 2015). These functions provide crucial benefits to both natural ecosystems and human societies, supporting environmental stability and resource availability (Smreczak et al., 2017).

The concept of soil functions gained scientific attention in the 1970s, when soil began to be recognized not only as a substrate for vegetation, but as a key ecological regulator (Blum, 2005; Smreczak et al., 2017). Soil plays a fundamental role in supporting plant growth by acting as a reservoir of nutrients and water while providing structural support for root systems. It also contributes to food and biomass production, ensuring agricultural sustainability and enabling ecosystems to function effectively. In addition, soil is essential for maintaining biodiversity, as it serves as a habitat for a vast number of plants, animals and microorganisms, many of which play a crucial role in nutrient cycling and ecosystem resilience. Soil also functions as a natural filter, regulating the transport of carbon, nutrients, and pollutants, and influencing global biogeochemical cycles. One of its most important environmental functions is carbon sequestration, which helps mitigate climate change by reducing atmospheric CO₂ levels. Beyond its ecological roles, soil provides valuable natural resources such as minerals and raw materials used in construction. It also serves as a repository of geological and archaeological heritage, preserving evidence of both Earth’s history and human civilization (Blum, 2005; Dominati et al., 2010; Gruszczyński, 2014; Morel et al., 2015; Greiner et al., 2017).

International organizations have also developed complementary frameworks for classifying soil functions, reflecting differences in priorities and policy objectives. Figure 2 compares selected soil functions as defined by the Food and Agriculture Organization of the United Nations (2015) and the European Commission (Stolte et al., 2016). According to the FAO, soil plays a critical role in providing food, fiber and fuel production, organic carbon sequestration, water purification and pollution reduction, climate regulation, and nutrient cycling. It also contributes to biodiversity conservation, flood mitigation, and the supply of raw materials for industry and infrastructure. Additionally, soil serves as a source of medicinal substances and genetic resources, reinforcing its significance in environmental and human health (Food and Agriculture Organization of the United Nations, 2015).

Figure 2

Figure 2. Comparison of soil function classifications as defined by the Food and Agriculture Organization of the United Nations and the European Commission. Own compilation based on institutional reports (Food and Agriculture Organization of the United Nations, 2015; Stolte et al., 2016).

Similarly, the European Commission emphasizes the importance of soils for biomass production, nutrient and water storage and filtration, carbon sequestration, and the maintenance of biodiversity at genetic, species, and habitat levels. Moreover, it highlights the role of soil as a repository of geological and archaeological heritage (Stolte et al., 2016).

These functional frameworks remain widely cited. However, more recent reports have expanded the perspective on soil roles, particularly in urban contexts. The 2020 report by the Intergovernmental Technical Panel on Soils (ITPS) highlights the importance of soil biodiversity for maintaining ecosystem resilience, and emphasizes the potential of urban soils for carbon sequestration, nutrient cycling, and supporting microbial life in anthropogenic environments (Food and Agriculture Organization of the United Nations, 2015). In recent years, remote sensing and Earth observation tools have been increasingly used to monitor and evaluate the spatial variability of soil functions, particularly in urban and peri-urban settings. Spectral indices, thermal imaging, and LiDAR can support assessments of soil sealing, erosion susceptibility, vegetation-soil interactions, and topographic structure–all of which are crucial for understanding the soil’s capacity to deliver ecosystem services.

3.2 Ecosystem services

Ecosystems are complex and dynamic systems composed of plants, animals and other environmental components that interact with one another, forming interdependent relationships (Smreczak et al., 2017). They possess self-organizing mechanisms that ensure stability, resilience, and functional integrity within landscapes (Solon, 2008). Ecosystem functions are the ecological processes that result in the provision of ecosystem services, and any changes in these functions directly affect the potential supply of services (Meulen et al., 2018). Ecosystem services refer to the benefits derived from natural processes and components, which directly or indirectly fulfil human needs (de Groot et al., 2002; Boyd and Banzhaf, 2007; Guerry et al., 2015). However, these benefits are only realized through human interaction with the environment, as ecosystems cannot generate services independently of human society (Costanza et al., 2014). To keep terminology consistent in the remainder of the manuscript, we distinguish soil functions from ecosystem services and apply a simple cascade throughout the review: soil properties and condition shape soil functions, and these functions provide the biophysical basis for ecosystem service outcomes. Table 1 summarizes how the key soil functions discussed in Section 3.1 relate to the four ecosystem service categories used here. This mapping is used later to organize the evidence on assessment approaches, including remote sensing indicators and valuation methods.

Table 1

Table 1. Typology of soil ecosystem services by function and service category.

For the purpose of this review, ecosystem services are grouped into four categories (Figure 3):

• regulating services,

• provisioning services,

• habitat services,

• cultural services.

Figure 3

Figure 3. Conceptual link between soil functions and ecosystem service categories in urban areas. The figure summarizes the framework used in this review to connect soil properties and condition with soil functions and, in turn, with ecosystem service outcomes. Services are grouped into provisioning, regulating, habitat and cultural categories. In the manuscript, this distinction is used to avoid treating “functions” and “services” as interchangeable terms and to clarify how specific soil functions underpin the service categories discussed in Sections 3,4.

Soils are closely tied to all four categories, although their contribution is often overlooked due to the challenges of monitoring underground processes. Remote sensing and spatial modelling have emerged as key methods to link aboveground vegetation patterns, land use, and surface sealing with belowground ecosystem service provision (Mulder et al., 2011; Araya et al., 2021). Vegetation indices such as the Normalized Difference Vegetation Index (NDVI) and the Enhanced Vegetation Index (EVI), as well as thermal imagery and LiDAR data, and classification of land cover allow indirect assessment of regulating and cultural services in urban landscapes.

3.2.1 Regulating services

Natural and semi natural ecosystems play a vital role in regulating fundamental ecological and life-supporting processes, such as climate stabilization, water purification, and air quality maintenance (de Groot et al., 2002). These regulating services allow humans to live in a stable, healthy and resilient environment (Dominati et al., 2010). Ecosystems contribute to air purification by filtering pollutants and preventing the spread of certain airborne diseases. They help maintain a favorable climate at local and global scales, which is critical for human health, agriculture, and economic stability (de Groot et al., 2002). Additionally, ecosystems serve as natural buffers against climatic extremes, reducing the impacts of storms, heat waves and floods through the absorption and dissipation of energy by vegetation (Gómez-Baggethun et al., 2013a) Soil plays a significant role in climate regulation by absorbing and storing atmospheric carbon, thereby mitigating the effects of greenhouse gas emissions.

The regulation of water dynamics ensures stable flow levels in rivers and canals, supporting irrigation and transportation networks. Soil plays an important part in nature’s water cycle. Its properties such as texture, structure, organic matter content, groundwater depth, water-holding capacity and bedrock perform an important role in water regulation. These properties, alongside vegetation and slope, determine the amount of rainfall infiltration and surface runoff, which in turn determines groundwater levels, peak flows, flood risk, erosion rates and the amount of sediment in flood waters (Keesstra et al., 2012; Meulen et al., 2018).

Another crucial function of regulating services is nutrient cycling, which plays a pivotal role in maintaining soil health, climate stability, and agricultural productivity. Ecosystems also regulate organic and inorganic waste decomposition, breaking down materials through dilution, assimilation, and chemical transformation (de Groot et al., 2002; Gómez-Beggethun et al., 2013b). They can also reduce atmospheric pollutants generated by transport and industrial activities. A significant role is played here by soil in both ways, directly by collecting contaminants from the air and indirectly as a basis for plants, which also have a positive impact on atmospheric quality. Coniferous trees are better suited to air filtration because the needles have a larger total surface area and they do not shed their needles in winter, when pollution is the greatest. On the other hand deciduous trees absorb gases better. The often-cited estimate that 1 hectare of mixed forest can remove up to 15 tons of particulate matter annually originates from urban ecology studies of the 1990s (Bolund and Hunhammar, 1999). More recent research, however, indicates that pollutant removal rates vary widely—from less than 1 kg/m² up to more than 5 kg/m² per year—depending on vegetation structure, species composition, surface properties, and local meteorological conditions (Janhäll, 2015). This suggests that the 15 t/ha value should be treated as indicative and context-dependent, and that future assessments should rely on locally calibrated models.

Soil and vegetation have the ability to reduce noise by absorbing, reflecting, and refracting sound waves, which has a significant impact on people’s quality of life (Gómez-beggethun et al., 2013b; Meulen et al., 2018). Dense shrubs and low trees have the greatest reduction effect because the branches that absorb sound are at human height. The design and shape of the greenbelt is of major significance. The density, height, length and width of tree belts are more important for noise reduction than the size of the leaves and the nature of the branches (Fang and Ling, 2003). Additionally, biological control is considered one of the regulating services. Evolution has shaped numerous interactions leading to fairly stable communities that naturally limit the spread of pests and diseases. Natural ecosystems control more than 95% of all potential crop pests and human disease transmitters (de Groot et al., 2002). Microorganisms play a special role in disease and pest control (Meulen et al., 2018).

3.2.2 Provisioning services

Provisioning services are the tangible goods that ecosystems supply, such as food, medicinal resources, raw materials, energy sources, and genetic material.

One of the most critical provisioning services is food production, which relies on crops, livestock, fisheries, and wild plant harvesting. A significant portion of global food supply also comes from wild plants, game, and aquaculture products, including fish, poultry, forest fruits, and mushrooms (de Groot et al., 2002). In urban environments, food production is limited, making cities heavily dependent on the surrounding rural areas (Gómez-Beggethun et al., 2013b). Another vital provisioning service is fresh water. Its quality and availability, including both surface and groundwater, are regulated by soil properties (Keesstra et al., 2012; Meulen et al., 2018).

Additionally ecosystems provide a variety of natural raw materials that are harvested for industrial purposes, including renewable biotic resources (wood, fiber), biochemical compounds (latex, rubber, oils, waxes, tannins, dyes), and various sources of energy (organic matter and wood). Many of these materials are derived from organic and mineral soil (de Groot et al., 2002; Comerford et al., 2013; Gómez-Beggethun et al., 2013b). Soil acts as the basis for infrastructure and buildings, which is strongly influenced by soil properties such as texture, structure and geohydrological conditions (Meulen et al., 2018).

Ecosystems are sources of many genetic resources. Many important crops need support from their wild relatives, especially in maintaining productivity or in the development of particular traits (for example, taste, resistance to pests and diseases, adaptation to environmental conditions). The soil environment produces chemical compounds that are used in the production of medicines, pharmaceuticals and cosmetics (de Groot et al., 2002; Comerford et al., 2013; Meulen et al., 2018).

3.2.3 Habitat services

Habitat services denote the role of ecosystems in providing living space and breeding grounds for countless species, thereby ensuring biodiversity and genetic diversity. Natural landscapes act as gene reservoirs, supporting the evolutionary processes necessary for species adaptation. Some habitats, such as wetlands, coral reefs, and forests, serve as nursery areas for young individuals before they migrate to other environments. Unfortunately, these critical ecosystems are often undervalued and frequently converted to more economic use, leading to disastrous ecological consequences (de Groot et al., 2002).

Soil has a fundamental role in habitat services, providing a home for microorganisms, invertebrates, and burrowing species. Soil is estimated to host between one-quarter and one-third of all living organisms on Earth, yet only 1% of soil microorganisms have been identified (Breure et al., 2012). In just 1 g of soil, more species of different organisms can be found than there are mammal species on the entire planet (Comerford et al., 2013). Soil organisms drive nutrient cycling and water purification, regulate the dynamics and structure of soil organic matter, and are involved in carbon sequestration and greenhouse gas emissions (Breure et al., 2012; Meulen et al., 2018).

3.2.4 Cultural services

Cultural services arise from the ways in which natural environments contribute to human wellbeing, through aesthetic value, recreation, spiritual significance and educational inspiration. They are often overlooked in scientific research, despite their significant influence on human culture, creativity, and psychological wellbeing (de Groot et al., 2002).

Nature has long served as a source of artistic inspiration, influencing literature, visual arts, music, fashion, and architecture. Many religions also incorporate sacred natural sites, such as forests, rivers, or mountains, which hold deep spiritual significance (de Groot et al., 2002). Soil, as an integral part of the landscape, plays an essential role in spiritual and cultural practices, serving as a place for burial, religious rituals, and historical preservation (Meulen et al., 2018).

One of the most significant cultural ecosystem services is the availability of natural spaces for rest, recreation, and physical activity (de Groot et al., 2002). This is especially relevant in urban environments, where green spaces, lakes, and forests enhance quality of life and reduce stress levels (Chiesura, 2004; Gómez-Baggethun et al., 2013b). Blue-green infrastructure increases property values and enhances social cohesion.

Ecosystems serve as natural laboratories, supporting environmental education, research, and long-term ecological monitoring. Natural areas are an important reference for monitoring environmental change. Soil is the site of much scientific research to understand the Earth’s history and nature (de Groot et al., 2002; Meulen et al., 2018).

4 Assessment of ecosystem services

4.1 Valuation of ecosystem services

Ecosystem services are difficult to value because most of them are not easily quantifiable, particularly cultural services. Different valuation approaches should be employed depending on the specific type of ecosystem service being assessed. Accurate and comprehensive valuation of ecosystem services can significantly support sustainable urban planning and effective development management strategies (Sudra, 2015).

The value of ecosystem services can be classified into ecological, socio-cultural and economic categories. Ecological values are determined by the integrity of regulating and habitat functions of the ecosystems and are assessed according to parameters such as complexity, diversity and rarity. Socio-cultural values play an essential role in identifying key environmental functions, such as, human physical and mental health, educational opportunities, cultural diversity and spiritual wellbeing. Natural systems are a source of intangible wellbeing and are fundamental to the development of sustainable societies (Dominati et al., 2010).

Spanish researchers conducted a study in Andalusia in the Nacimiento and Adra river catchment area, where they interviewed both residents and tourists about perceptions of ecosystem services. The research enabled the classification of services according to their socio-cultural valuation. The first category identified was essential services, encompassing primarily provisioning services (e.g., agriculture, animal farming, timber, fresh water) but also regulating services (e.g., erosion control, air and water quality) and aesthetic values. The second group was termed “important but insensitive services”, which included rural and natural tourism, peacefulness, relaxation and renewable energy. In contrast, “sensitive but less important” services such as local ecological knowledge, species habitat, beekeeping, soil fertility and fiber production were perceived as valuable but not critical. Lastly, the “less important services” category included certain cultural services and some provisioning services. The study highlighted that the valuation of ecosystem services varies significantly among different stakeholders, such as local and environmental professionals, service-dependent residents, residents directly independent of ecosystem services and tourists (Iniesta-Arandia et al., 2014). The value of an ecosystem service is not static. It depends on the service’s availability, the demand for it, and the existence of technological or natural alternatives. Moreover, perceived value can vary across spatial and temporal scales and among different stakeholder groups (Meulen et al., 2018). When monetary valuation is not feasible, ecosystem services can be assessed through their impact on human health and wellbeing (Guerry et al., 2015). Recent global assessments have underscored the dynamic nature of ecosystem service values particularly in the context of accelerating environmental degradation. This was also emphasized by the IPBES Global Assessment (2019), which called for more pluralistic approaches to valuing nature beyond monetary metrics. While the seminal work of Guerry et al. (2015) remains highly relevant in highlighting the integration of natural capital into decision-making frameworks, more recent updates suggest notable shifts in the estimated global value of ecosystem services over time. For instance, Costanza et al. (2017) demonstrated that the global loss of ecosystem services between 1997 and 2011 attributed largely to land use change, amounted to an estimated USD 4.3–20.2 trillion annually. This represents a significant increase compared to earlier estimates from 2014 (Costanza et al., 2014), reflecting both improved valuation methods and escalating ecosystem degradation. These findings emphasize the importance of incorporating temporal trends and scenario-based modelling in ecosystem service valuation, particularly in rapidly urbanizing regions where land transformation is most pronounced. Including updated estimates strengthens the evidence base for integrating ecosystem service considerations into economic and policy frameworks.

Monetary valuation is possible for goods and services that are exchanged on the market and have a clearly defined price, such as peat sold in garden shops, silica used in ceramics or various soil organisms (e.g., ants, termites, earthworms) consumed as a protein source. It is critical to distinguish use value from market value, as services with high use value may have little or no market value; an example is the availability of clean drinking water (Baveye et al., 2016).

As many ecosystem services lack direct market prices, a range of economic valuation methods have been developed to support environmental decision-making. The valuation of services that cannot be assigned a direct market price is inherently more complicated. In such cases, alternative economic valuation methods are employed, including:

• estimating the monetary value of services based on their contribution to the production of market goods,

• assessing the cost of avoided damages that would result from service loss,

• calculating the cost of replacement or substitute services, such as flood control provided by wetlands or temperature regulation through urban tree planting,

• evaluating the impact of ecosystem services on market prices, for example, the higher property values of apartments located near parks,

• using the travel cost method to estimate the value of ecosystems based on the amount of money people spend to access them, particularly relevant for cultural services rather than soil services,

• conducting contingent valuation studies that involve surveying individuals to determine their willingness to pay for the maintenance of specific ecosystem services under hypothetical scenarios,

• implementing group valuation methods, which are based on democratic deliberations and public debates to assess collective preferences (Chee, 2004; Dominati et al., 2010; Baveye et al., 2016).

Ecosystem services are considered public goods, as no one can claim exclusive ownership or prevent others from benefiting from them. As a result, managing these services sustainably is particularly challenging. Consequently, it is difficult to hold individuals or institutions accountable for activities that negatively impact these services. Ecosystem services are often overlooked in traditional economic analyses and policy decision-making processes. Effective valuation of ecosystem services can foster a better understanding of their importance, illustrate the distribution of their benefits, facilitate the equitable distribution of management costs, and promote the development of innovative institutional and market-based instruments to support sustainable ecosystem management (Chee, 2004). Additionally, the perceived value of ecosystem services varies significantly across spatial and temporal scales, and among stakeholder groups with different dependencies, experiences, and priorities (Iniesta-Arandia et al., 2014; Meulen et al., 2018). Figure 4 illustrates an integrative framework for valuing ecosystem services using ecological, socio-cultural, and economic approaches. It highlights the potential of remote sensing data to inform each of these domains—ranging from biophysical indicators and spatial modelling to participatory mapping and economic valuation tools. By linking Earth observation technologies with multiple valuation dimensions, the framework supports comprehensive assessments of ecosystem service benefits at landscape scale.

Figure 4

Figure 4. Conceptual framework for the typology of ecosystem service valuation methods, integrating remote sensing data with ecological, socio-cultural, and economic approaches. Own study based on Chee (2004), Gómez-Baggethun et al. (2013b) and recent remote sensing applications.

4.2 Remote sensing applications for soil ecosystem services

Monitoring soil ecosystem services requires a deep understanding of spatial and temporal changes occurring both above and below the surface. In recent years, remote sensing has become an indispensable tool for assessing soil health, land cover dynamics, and ecosystem functionality at various scales. The integration of satellite imagery, aerial observations, and drone-based surveys offers an unprecedented opportunity to detect soil degradation processes, quantify ecosystem services, and inform sustainable land management.

Satellite missions such as Sentinel-2 and Landsat-8 provide high-resolution multispectral imagery that enables the monitoring of soil properties, vegetation cover, and land-use changes. Indices derived from spectral data, including the Normalized Difference Vegetation Index (NDVI) and the Normalized Difference Built-up Index (NDBI), have been successfully used to assess vegetation productivity, detect impervious surfaces, and evaluate urban soil sealing (Small and Milesi, 2013; Greiner et al., 2017). NDVI, in particular, has become a standard measure for estimating primary productivity and indirectly inferring soil quality through vegetation health. Table 2 provides a comparative overview of key remote sensing technologies used in urban soil monitoring. It summarizes their spatial and spectral resolutions, primary applications related to soil ecosystem services, and the main advantages and limitations of each approach. Working with multiyear satellite image archives also enables change detection approaches that track land cover and surface dynamics over time, rather than relying on a single snapshot (Michałowska et al., 2016).

Table 2

Table 2. Comparison of remote sensing technologies in urban soil monitoring.

In the context of ecosystem service assessment, it is helpful to distinguish between variables that remote sensing observes directly and services that can only be inferred indirectly. Land cover, impervious surface extent, vegetation structure and land surface temperature are directly observable and can serve as robust indicators for selected regulating and cultural services, for example, sealing related runoff risk and local cooling demand. By contrast, services dominated by belowground processes, including soil carbon dynamics, nutrient cycling, pollutant retention and most aspects of soil biodiversity, cannot be measured from imagery and are typically estimated through proxies and models that combine remote sensing with ancillary data. These inferences are scale sensitive: mixed pixels, shadowing, seasonality and heterogeneous management can weaken proxy validity in urban mosaics. For this reason, remotely sensed indicators are best treated as screening and monitoring tools and should be complemented by field measurements for calibration and validation whenever results are intended to inform planning or valuation.

Recent advances in hyperspectral imaging and Light Detection and Ranging (LiDAR) technologies have further enhanced the capacity to map soil characteristics, such as organic matter content, salinity, and compaction levels (Chabrillat et al., 2019). These technologies improve the spatial representation of soil properties and surface structure that underpin key soil functions relevant to ecosystem services, including carbon storage potential, water regulation and nutrient cycling, but field measurements remain necessary for validation.

Drones equipped with multispectral and thermal sensors offer complementary, fine-scale data, particularly valuable for urban and peri-urban areas where soil sealing and fragmentation patterns are highly heterogeneous. UAV-based surveys can detect subtle changes in vegetation vigor, surface roughness, and soil moisture, which are crucial indicators of soil ecosystem service provision in human-modified landscapes (Greiner et al., 2017).

Remote sensing thus plays a key role in bridging the gap between field-based soil studies and regional or global assessments of ecosystem services. Figure 5 presents a proposed workflow integrating various remote sensing tools (satellite indices, UAV imagery, and LiDAR) for assessing soil ecosystem services in urban environments. The diagram illustrates a stepwise progression from data acquisition to mapping and interpretation of ecosystem services. Such a workflow supports effective monitoring and early warning of soil degradation, and provides critical inputs to spatial models that quantify and map soil functions. It also emphasizes that combining remote sensing with field observations is essential to ensure data accuracy and interpretation reliability (Gómez-Baggethun and Barton, 2013a). The following section describes the specific use of remote sensing methods in assessing and monitoring specific ecosystem services in urban areas.

Figure 5

Figure 5. Remote sensing workflow for soil ecosystem service assessment (own study).

5 Threats to soil ecosystem services and potential solutions

Urbanization is one of the major threats to the environment and soil cover. This phenomenon is associated with the migration of people from the rural areas to cities and steady increase in the proportion of urban residents relative to the total population, combined with the spatial expansion of urban areas through intensification of development (Łuczyszczyn and Łuczyszczyn, 2018) and redefinition of administrative boundaries. Today, urban areas are growing more rapidly than the urban population itself (Seto et al., 2013). Historically, the first human settlements were established on fertile soils located on plain areas, in river valleys, or near mineral deposits, areas favorable for cultivation and resource extraction. As a consequence, soils with high agricultural suitability have been, and continue to be, most frequently degraded by urban development (Famielec et al., 2007).

The main pressures imposed by urbanization on soil functions and the resulting impacts on ecosystem services are synthesized in Table 3, providing a structured overview of cause-effect relationships.

Table 3

Table 3. Classification of urbanization threats to soil functions and ecosystem services.

Urbanization is a complex and dynamic process that occurs across multiple spatial and temporal scales. In ancient times, humans migrated in search of water and natural resources. Following the invention of agriculture, permanent settlements were established in many regions of the world. Although urban communities thrived through specialization in goods production, they remained dependent on their rural surroundings for the exchange of goods and services. For many centuries, densely packed dwellings and central institutions surrounded by residential extensions were characteristic of cities in the Middle East, Asia and Europe (Elmqvist et al., 2013).

Currently, urban areas are growing at approximately twice the rate of urban population. In 1950, 29% of the global population resided in cities; by 2050, it is estimated that this figure will reach approximately 70% with populations primarily inhabiting small and medium-sized cities (Seto et al., 2013). Continued urban expansion is therefore an inevitable challenge. One of the most damaging impacts of urbanization on soil is soil sealing, the process whereby the topsoil or entire soil profile is destroyed or covered with impermeable materials, such as buildings and infrastructure. Soil sealing represents the most intensive and often irreversible form of land transformation (Stolte et al., 2016). Urbanization exerts pressure on the environment by increasing the demand for natural resources required for construction. Moreover, changing lifestyles and consumption patterns associated with urban living contribute to greater demands for energy and water resources (Seto et al., 2013). Sealed soils lose their ability to provide essential ecosystem services required for maintaining habitats suitable for humans, animals, and plants. Given the inevitability of urban expansion, it is crucial to implement strategies aimed at minimizing soil degradation and to promote environmentally sustainable urban planning.

Urban soils encompass the soils of urban, industrial, communication, mining and military areas (O’Riordan et al., 2021). These soils have been heavily modified by human activities including mixing, importing and exporting materials, and are often characterized by contamination, compaction, sealing, and the disruption, or alteration of natural soil horizons (Li et al., 2018; O’Riordan et al., 2021). Urban soils are commonly found in parks, along roads, in sports fields, along urban rivers, in suburban zones, near landfills and mining areas, and adjacent to industrial and communication facilities (Guilland et al., 2018; Li et al., 2018). They are typically distinguished by the presence of anthropogenic artefacts, deep material mixing and the formation of artificial organic or mineral layers, many of which are impermeable to water and plant roots (Polskie Towarzystwo Gleboznawcze, 2019). Importantly, urban soils have the potential to store significant amounts of carbon, particularly in arid climates (Pouyat et al., 2006).

As a vital component of the urban ecosystem, urban soils must be managed effectively to ensure soil quality sufficient to sustain plant and animal productivity, maintain or enhance water and air quality, and support human health and habitats (Li et al., 2018). Unsealed urban soils provide a range of ecosystem services critical to city environments, particularly regulating and cultural services (Sudra, 2015). Regulating services include mitigating flooding risks, buffering the urban heat island (UHI) effect, capturing pollutants, and supporting nutrient cycling and carbon storage. Additionally, they support diverse soil biota (Setälä et al., 2014). Cultural services offered by urban soils include providing physical support for urban infrastructure and facilitating access to green spaces, which are crucial for enhancing the physical and mental wellbeing of urban residents (O’Riordan et al., 2021).

5.1 Threats to climate conditions in urban areas

Land use changes associated with urban development significantly influence temperature and precipitation patterns in and around urban areas. Built-up areas alter the exchange of heat, water, gases and aerosols between the land surface and the atmosphere (Seto et al., 2013). In the short term, changes primarily occur in the upper layer of the soil, as a heat transfer within soil is a slow process. Temperature anomalies in the surface layer are subsequently released into the atmosphere, while only persistent anomalies are transmitted into deeper soil layers, where they can disrupt soil chemical processes. The temperature of sealed soils depends on the albedo, emissivity and thermal properties of the sealing material (Scalenghe and Marsan, 2009).

Urbanized areas experience increased surface runoff, which reduces evaporation and diminishes natural cooling processes. This leads to the formation of ‘urban heat island’ (UHI) – a meteorological phenomenon characterized by higher air temperatures in urban areas compared to adjacent rural areas (Trząski, 2015; Stolte et al., 2016). UHI development is influenced by the shape, size and spatial arrangement of buildings (Seto et al., 2013). The phenomenon was first documented in the early 1900s, when a temperature difference of 2.5 °C between central London and surrounding areas was observed (Soltani and Sharifi, 2017). In the United States, diurnal temperature variations in large cities can reach 10 °C–12 °C, while in Poland the differences range from 5 °C to 8 °C (Fortuniak, 2003). The urban/rural temperature differential is typically greater during clear, calm nights and in winter months (Soltani and Sharifi, 2017). Throughout the year, the largest temperature disparities occur during spring and summer, with a notable decrease during winter (Adamczyk et al., 2008). The type of urban development also plays an important role in shaping the urban climate. Areas characterized by dispersed and tall buildings exhibit climatic conditions similar to those of rural areas and promote greater thermal comfort during summer. The most favorable conditions for residents occur in districts featuring diverse buildings, structures and abundant green spaces (Osińska-Skotak, 2002).

The urban heat island effect negatively impacts both the physical and mental health of city dwellers, particularly the elderly and chronically ill. Instances of thermal discomfort and heat stress are becoming more frequent (Nidzgorska-Lencewicz and Mąkosza, 2016). Intensifying urban heat leads to economic losses (e.g., higher electricity demand for air conditioning, decreased tourism) and non-economic losses (e.g., decreased resident wellbeing). Heat stress persisting for more than three-quarters of a day should be classified as a crisis situation (Sikora, 2008).

Furthermore, sealed surfaces exacerbate temperature extremes, and tall buildings disrupt air circulation (Gerst et al., 2011). Urbanization also contributes to an increase in extreme weather events such as droughts, storms, and intense rainfalls (Solecki and Marcotullio, 2013). Local urban climate is further affected by aerosols, air pollutants and carbon dioxide emissions from transportation and industry. Pollutants can scatter, reflect or absorb solar radiation, with aerosols exhibiting either a cooling effect (e.g., sulphates) or a warming effect (e.g., black carbon) (Seto et al., 2013).

Urbanization also influences precipitation patterns, creating the so-called ‘urban rain effect’ – characterized by increased total rainfall, higher frequency of rainy days, longer precipitation duration, and greater incidence of hail, thunderstorms, and heavy rainfalls. Factors contributing to the urban rain effect include thermal convection, strong vertical air currents, the urban heat island phenomenon, the presence of condensation nuclei from pollutants, and elevated atmospheric water vapor content. While increased rainfall can improve air quality, it simultaneously poses challenges such as groundwater contamination and transport disruptions (Nidzgorska-Lencewicz and Mąkosza, 2016).

Urbanization also affects the carbon cycle. Urban soils can accumulate greater amounts of carbon compared to natural soils (O’Riordan et al., 2021). Residential lawn soils often exhibit higher carbon densities than nearby forest soils (Pouyat et al., 2006). Individual unsealed land parcels in cities can store up to 14 kg/m2 of carbon, while the city-wide average is around 6 kg/m2 (Setälä et al., 2014). More than half of the organic carbon is concentrated in the topsoil layer, which is permanently lost when land is sealed by buildings and infrastructure (European Commission, 2012; Stolte et al., 2016).

Modern satellite observations significantly support the analysis of the urban heat island phenomenon. Thermal infrared data from satellites (e.g., Landsat 8 thermal bands and NASA’s ECOSTRESS instrument on the International Space Station) enable the creation of detailed maps of surface temperature distributions in cities. For example, thermal imagery clearly records higher roof and asphalt temperatures in city centers compared to green areas, allowing identification of UHI “hotspots” (Small and Milesi, 2013; Greiner et al., 2017; Głowienka et al., 2025a). Long-term archives (e.g., Landsat since the 1980s) allow researchers to track UHI changes over time and assess the impact of urbanization on local climate. The growing use of platforms such as Google Earth Engine has significantly improved the processing of large satellite image datasets for ongoing monitoring of urban thermal conditions (Kara et al., 2025). Data from Landsat, ECOSTRESS, MODIS, and Sentinel programs are freely and regularly available, allowing for the construction of dense time series and comparisons across cities (Guha et al., 2018; Głowienka et al., 2025b). Studies show a strong negative correlation between NDVI and surface temperature and a positive correlation between NDBI and LST (Guha et al., 2018; Kamiński et al., 2025; Kara et al., 2025). This means that urban areas with limited greenery tend to exhibit higher surface temperatures, while vegetated zones remain relatively cooler. Information on temperature distribution and its relationship to land cover provides a valuable basis for spatial planning by identifying areas for adaptation measures, such as expanding green areas or using cool surface materials. The use of remote methods enables the identification of urban zones most at risk from UHI effects and monitoring the effectiveness of implemented mitigation strategies (Kamiński et al., 2025). The ability to integrate satellite data with ground observations and urban climate models makes remote sensing an indispensable tool for assessing climate risks in cities and planning their resilience to extreme weather events (Głowienka et al., 2025a, Głowienka et al., 2025b).

Remote sensing also facilitates monitoring of urban water resources–for instance, radar satellites like Sentinel-1 and the SMAP radiometer provide soil moisture data, which indirectly indicates infiltration efficiency and the land surface’s cooling potential. Such remote observations are invaluable for city planners, highlighting where interventions (e.g., planting greenery or using permeable surfaces) are most urgently needed to alleviate extreme urban temperatures (Zhou et al., 2019; Jha et al., 2024; Araya et al., 2021; Zhao et al., 2025).

5.2 Threats to biodiversity

Although urbanized areas occupy only a small fraction of the Earth’s surface, they are responsible for a disproportionately large share of anthropogenic changes in the biosphere (Seto et al., 2013). Urban areas are deeply interconnected with their surrounding ecological landscapes. Ecosystem services consumed within cities are often generated by ecosystems located far beyond urban boundaries, sometimes on the other side of the world (Gómez-Baggethun et al., 2013b). Historically, cities were frequently established in areas of high species richness, such as islands, coastal zones, and major river valleys, which are now heavily urbanized (Seto et al., 2013). Urban expansion often leads to soil sealing on fertile lands, significantly impacting food security (Gruszczyński, 2014). Biodiversity is an essential component of many ecosystem services, yet it is fundamentally and irreversibly altered by anthropogenic activities.

The direct impact of urbanization on biodiversity includes changes in land cover associated with urban expansion, which increasingly encroach upon valuable and protected natural areas (McDonald et al., 2013). The number of endangered species is steadily increasing, and the biological uniqueness of ecosystems is threatened by the displacement of native species by invasive foreign ones. Over recent decades, there has been a marked decline in the diversity of wild bees and bumblebees in Europe and the United States. These organisms play a critical ecological role as pollinators of wildflowers and agricultural crops (Ahrné et al., 2009). Urban development fragments ecological corridors and habitats, disrupting the migration routes of plants and animals (Scalenghe, and Marsan, 2009; Stolte et al., 2016).

Soil ecosystems are also critically impacted. For many insect species, soil serves as a breeding, nesting or foraging habitat (European Commission, 2021; Stolte et al., 2016). Urbanization results in a decline in microbial biodiversity, which is increasingly linked to higher incidence of allergies and asthma among urban populations. Modern building design has led to more sterile living environments and diminished natural exposure to environmental microbiota, thereby weakening immunological resilience (Li et al., 2018). Microorganisms perform vital ecological functions, including organic matter decomposition, nutrient cycling, and carbon dioxide absorption and storage. Urbanization threatens both belowground and aboveground biodiversity. Notably, approximately one-quarter of the Earth’s terrestrial biodiversity is estimated to reside in soils, including microorganisms, invertebrates, and soil-dwelling fauna (European Commission, 2012; Stolte et al., 2016; Food and Agriculture Organization of the United Nations, 2015). Soil sealing interrupts the biological and chemical cycles of terrestrial organisms, trapping them within the soil matrix and leading to a breakdown of ecological processes essential for ecosystem stability (Gruszczyński, 2014).

Monitoring urban biodiversity and habitat fragmentation increasingly relies on remote sensing technologies. High-resolution satellite imagery (e.g., Sentinel-2, WorldView) and UAV (drone) mapping allow detection of land cover changes, vegetation diversity, and the fragmentation of ecological corridors. For instance, using land-cover indices derived from Sentinel-2 data, researchers can map urban green infrastructure and identify key areas that maintain habitat continuity (such as parks, woodlands, and river valleys) (Rocchini et al., 2018; Finizio et al., 2024). Spatial analyses based on satellite images help assess where urbanization most severely disrupts ecosystem connectivity, which in turn supports the planning of new green spaces or wildlife corridors. In addition, drones can be used to inventory urban vegetation with high precision—for example, counting individual trees and evaluating their health via vegetation indices. LiDAR remote sensing data, providing 3D models of vegetation structure, offer insights into habitat complexity (e.g., variability in tree height), which correlates with species richness of birds, insects, and other taxa. This suite of remote monitoring tools aids biodiversity conservation in cities by pinpointing areas in need of ecological restoration and by enabling the tracking of intervention outcomes (such as tree planting or creation of wildflower meadows) over time (Rocchini et al., 2018; Randin et al., 2020; Finizio et al., 2024).

5.3 Threats to water cycle

Urban development significantly disrupts the natural water cycle. Soil sealing reduces the water retention capacity of soils and substantially increases surface runoff, which can double when impervious surfaces exceed 20% of the land area (O’Riordan et al., 2021). Estimates indicate that in sealed urban areas, up to 60% of rainfall becomes surface runoff, compared to only 5%–15% in vegetated zones. This pattern has been consistently documented in both earlier and more recent European assessments (Bolund and Hunhammar, 1999; European Commission, 2012; Stolte et al., 2016). As a result, sealed soil loses its ability to store water, leading to increased demands on storm water drainage systems and higher associated economic costs. In urbanized areas, infiltration rates decrease, lowering groundwater recharge and consequently reducing the availability of drinking water in cities, as well as limiting water accessibility for urban vegetation. The decline in groundwater levels deteriorates conditions for the growth and health of urban green spaces, which are critical for improving air quality by absorbing atmospheric pollutants and dust on their leaf surface (Stolte et al., 2016).

Additionally, reduced groundwater levels diminish the intensity of chemical reactions within soil processes (Scalenghe and Marsan, 2009). Urbanization also alters key factors affecting water quality, such as erosion processes, nutrient cycling and biogeochemical dynamics (McDonald et al., 2013). Even a minor increase in impervious surface area by as little as 2% within an urban catchment can significantly alter storm water chemistry, including elevated salinity and changes in pH levels (Scalenghe and Marsan, 2009). Increased pressure on water resources leads to deterioration of river basin ecosystems, particularly impacting water-related ecosystem services (European Commission, 2012; Gruszczyński, 2014; Stolte et al., 2016). Neighboring non-urbanized areas are also affected by sealing. Accelerated surface runoff increases the risk of flooding and soil erosion on adjacent unsealed lands. Pollutants accumulated on impervious surfaces are washed away during heavy rainfall events, subsequently infiltrating into neighboring soils and degrading their quality (Scalenghe and Marsan, 2009). Furthermore, urbanization demands large quantities of water for agriculture, industry and direct human consumption (e.g., cooking, cleaning, sewage disposal). In an arid climate, the availability of water resources becomes a particularly critical issue (McDonald et al., 2013).

The process of urban soil compaction—caused, for example, by heavy machinery traffic, transportation, or construction activities—represents a significant threat to the hydrological and biogeochemical functions of soil. Compaction results in the densification of soil particles and the loss of porosity, particularly macropores responsible for aeration and rapid water movement. Studies have demonstrated that even moderate levels of compaction can drastically reduce soil infiltration capacity—rainwater, instead of percolating into the ground, runs off the surface, thereby increasing the intensity and volume of surface runoff (Gregory et al., 2015). At the scale of urban catchments, this translates into a higher risk of flash flooding during heavy rainfall events, as compacted and sealed soils lack the capacity to absorb excess water. Reduced infiltration also leads to diminished groundwater recharge and exacerbates soil desiccation during dry periods, which impairs the growth conditions for urban vegetation.

Compaction affects not only water cycling but also soil aeration dynamics. The reduction in air-filled porosity limits oxygen diffusion into the soil profile; in extreme cases, the rate of oxygen consumption by soil biota exceeds the rate of supply, resulting in the formation of anoxic (oxygen-depleted) zones. Oxygen-deprived soils lose their capacity to support a wide range of ecosystem services associated with aerobic microbial activity. Under anoxic conditions, the composition of the soil microbiome shifts—anaerobic microorganisms such as sulphate reducers and methanogens become dominant (Nawaz et al., 2013). This transition leads to a significant decline in the efficiency of natural degradation of organic pollutants. Under aerobic conditions, many contaminants (e.g., petroleum hydrocarbons or organic compounds from wastewater) undergo biodegradation facilitated by bacteria and fungi. In contrast, in compacted and oxygen-limited soils, these degradation processes are slowed or proceed via alternative, less efficient anaerobic pathways (Gregory et al., 2015). Moreover, oxygen deficits favor the formation of reduced, often more toxic, contaminant intermediates and inhibit the oxidative immobilization of heavy metals. As a result, soil compaction in urban environments exacerbates contamination problems—both by increasing the transport of pollutants via runoff (due to restricted infiltration) and by impairing the soil’s natural filtering capacity. For these reasons, recent literature recommends that structural degradation of soil be explicitly included in assessments of urbanization impacts on soil systems, alongside commonly analyzed processes such as surface sealing and chemical pollution.

Remote sensing tools play a key role in monitoring urbanization-induced changes in the water cycle. Satellite land-cover maps (e.g., classifications from Landsat or Sentinel data) allow precise determination of what fraction of an urban catchment is impervious, providing essential input data for hydrological models that estimate surface runoff and flood risk (Shao et al., 2023). Studies demonstrate that satellite observations enable the production of high-resolution maps of impervious surfaces in cities, greatly improving our ability to assess urban runoff dynamics and flooding hotspots. Furthermore, satellite radar measurements (Sentinel-1) and passive microwave data (e.g., NASA’s SMAP mission) provide maps of soil moisture, which can be used to track how quickly urban soils dry out after rainfall and to identify areas of severe desiccation (for instance, due to soil compaction) (Araya et al., 2021). This information helps pinpoint zones with impaired infiltration that may require remediation (such as soil decompaction or aeration). Optical remote sensing is also applied to monitor water quality–for example, after intense rainfall events, satellite imagery from Sentinel-2 often reveals plumes of suspended sediment in urban rivers and reservoirs, indicating increased erosion and runoff of pollutants from developed areas (Kong et al., 2025). Such remote sensing data support urban storm water management by enabling assessment of the performance of blue-green infrastructure (e.g., rain gardens, green roofs, retention basins) in reducing surface runoff and enhancing infiltration (Ahmad and Hassan, 2024).

5.4 Contamination

Urbanization contributes to increased environmental pollution, primarily originating from transportation, coal combustion, industrial activities and waste disposal. Soil contamination is defined as the presence of pollutants in the soil at concentrations exceeding established safety thresholds, leading to the degradation or loss of one or more soil functions. Contamination affects both the soil itself and the broader ecosystem. Soil pollution negatively impacts biomass production, disrupts the carbon cycle, diminishes the soil filtering and buffering capacities, and poses serious risks to human health (Stolte et al., 2016). Human health is threatened through two principal pathways. The first involves direct contact with contaminated soil, such as during recreational activities in urban parks. The second pathway is indirect exposure through the soil-plant-human chain, when contaminated suburban fields or home gardens lead to the production of contaminated crops and food products (European Commission, 2012; Stolte et al., 2016; Li et al., 2018). Additionally, the widespread use of road salt poses a significant threat to soil quality, contributing to soil degradation, salinization, and disruption of soil microbial communities (O’Riordan et al., 2021).

New remote sensing techniques are increasingly enabling large-scale identification of soil contamination. Hyperspectral imaging provides very detailed spectral information, making it possible to detect anomalies associated with the presence of heavy metals or chemical pollutants in soil. For example, lead or petroleum contamination in soils can cause characteristic changes in reflectance in the near-infrared range, as observed in airborne spectral surveys (Chabrillat et al., 2019). Satellite missions such as the Italian PRISMA and the upcoming German EnMAP offer the ability to map soil properties like organic matter and mineral content, and potentially to identify contamination “hotspots” based on unique spectral signatures of specific compounds (Lovynska et al., 2024). Moreover, remote sensing can indirectly indicate soil contamination by analyzing vegetation conditions: areas where plants exhibit stress (unusually low NDVI or altered coloration) may point to toxic soil conditions (Dean et al., 2024; Lovynska et al., 2024). Moreover, deep learning applied to multi-temporal multispectral imagery enables spatial estimation of soil Pb (Tan et al., 2023). Finally, thermal and moisture data can help locate contaminated sites–for instance, salt-affected soils tend to dry out faster and heat up more than surrounding areas, which is detectable in thermal imagery. Although these methods require careful calibration and on-site validation, they represent a promising set of tools for early detection of soil contamination and for monitoring the effectiveness of soil remediation efforts (e.g., cleanup of brownfields or post-industrial lands) (Mangafić et al., 2025). Practical examples also show that freely available satellite imagery can support repeated monitoring of reclaimed postindustrial sites, which is directly relevant for cities dealing with brownfields and legacy land use (Hejmanowska et al., 2016).

5.5 Mitigation and planning strategies

To address the above threats, researchers and planners propose various mitigation strategies. Recent literature highlights the need for more robust integration of soil functions and ecosystem services into urban spatial planning frameworks (Teixeira da Silva et al., 2018). Traditionally, soil-related aspects have been underrepresented in urban policy, despite the pivotal role that soils play in sustaining urban sustainability, such as regulating water flows, sequestering carbon, and supporting biodiversity. The ecosystem services framework offers a useful instrument for translating soil-related ecological processes into a language relevant to policy-making and urban design, enabling the benefits provided by soils to be expressed in terms that are meaningful to planners and decision-makers (Orlova and Savin, 2024). However, reviews of current planning practices indicate that most urban plans still fail to incorporate explicit indicators of soil condition or ecosystem services, even when soils are nominally acknowledged as important assets. There is therefore a gap between knowledge about soil ecosystem services and their practical implementation. The development of simple and accessible indicators of soil ecosystem services is recommended to support urban planners and decision-makers who are not specialists in soil science (Teixeira da Silva et al., 2018). In practice, three barriers recur across cities. First, responsibility for soils is often fragmented between planning, environmental protection, transport and water management units, which makes it difficult to translate ecosystem service indicators into binding planning provisions. Second, legal instruments frequently focus on contamination control and do not explicitly require the protection of soil functions, so soil-related criteria enter decision-making late, during project permitting. Third, data limitations remain substantial: many municipalities lack harmonized soil maps, routine monitoring, or locally calibrated thresholds that would allow ecosystem service indicators to be used operationally. As a result, remote sensing products are increasingly used as screening tools, but they still require targeted field verification when they inform zoning decisions or investment priorities (Greiner et al., 2017; Teixeira da Silva et al., 2018).

The implementation of green infrastructure and nature-based solutions (NBS) is increasingly promoted as a means of integrating ecosystem services, including those provided by soils into urban planning. This approach involves the deliberate use of natural elements (soils, vegetation, and rainwater flows) to enhance urban resilience to environmental hazards and to improve the overall quality of the urban environment (Ahmad and Hassan, 2024). Green infrastructure improves the quality of unsealed urban soils (Celletti et al., 2025). The use of different types of NBS responds to various contemporary social challenges faced by urban residents. The most effective elements are those that use natural processes (e.g., natural areas with high or low vegetation or ponds) and those that use and mimic natural processes (such as green roofs, green walls, artificial wetlands, or planting pit systems). The least effective are elements that mimic natural processes (such as permeable pavements or rainwater tanks). An analysis conducted by Goličnik Marušić et al. (2023) shows that natural and semi-natural elements contribute to decreasing greenhouse gas emissions and lowering temperatures through carbon storage, shading, and lower heat emissions. In addition, they improve urban water cycle management by reducing surface runoff and purifying and storing water. They also contribute to enriching air quality by lessening pollution and producing oxygen. Artificial elements, on the other hand, are mainly engaged in urban water cycle management. However, implementing the NBS concept into planning practices requires appropriate policy frameworks and a recognition of the value of natural capital. Furthermore, it is important to consider different aspects, perspectives and stakeholders in order to develop urban areas with minimal loss of ecosystem services (Arkema et al., 2015). Accordingly, international initiatives such as the EU Green Infrastructure Strategy and the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services, 2019 and Ecosystem Services (IPBES) actively promote the inclusion of ecosystem services in spatial planning procedures and decision-making processes (Ahmad and Hassan, 2024; European Commission, 2025). As a result, urban planning is gradually shifting its emphasis from conventional grey infrastructure towards nature-based approaches, which are often more cost-effective and multifunctional. Nature-based solutions in urban areas should be understood as a network of solutions that improve the quality of the environment and life of residents, rather than as individual, separately implemented solutions. A single NBS does not cover all the natural processes needed in a given area, but the implementation of several NBS in various combinations can effectively address the identified social challenges in a city (Goličnik Marušić et al., 2023). In the long term, such strategies foster sustainable urban development while safeguarding natural capital and improving the quality of life for urban residents.

One example of a framework introducing NBS into spatial planning is the study by Raymond et al. (2017). Based on scientific documents and taking into account various social challenges related to cities, the authors created a 7-step process for implementing NBS benefit assessments in policy and spatial planning. The first step is to identify the problem and select and assess NBS. The next step is to develop an implementation strategy and implement nature-based solutions, followed by their scaling. Throughout the process, NBS are monitored and their additional benefits are evaluated. In addition, various stakeholders are involved throughout the process. The implementation of nature-based solutions addresses urban issues and challenges that are multidimensional and complex, which is why their selection and valuation requires the participation of a wide range of stakeholders, including multidisciplinary scientific teams, policymakers, and decision-makers (Raymond et al., 2017).

Remote sensing tools offer valuable spatial data that can significantly enhance ecosystem-based urban planning. They support repeated monitoring of key parameters related to ecosystem services, such as land cover distribution, surface temperature, and soil moisture, and facilitate the assessment of the performance of green infrastructure and nature-based solutions over time. Satellite and aerial imagery provide urban planners with measurable indicators, including the share of impervious surfaces, the extent of vegetated areas with biological activity, and the spatial intensity of urban heat island (UHI) effects. These data are instrumental in incorporating soil-related ecosystem services into urban models and GIS-based planning systems (Teixeira da Silva et al., 2018; Guo et al., 2022; Ahmad and Hassan, 2024; Orlova and Savin, 2024; European Commission, 2025). Integrating such geospatial information helps planners identify areas requiring intervention, evaluate the effects of implemented measures, and develop more sustainable and resilient urban environments.

In the mitigation literature, labels matter less than what is actually done on the ground. Measures deliver lasting benefits when soil condition and soil functioning are considered from the outset and then supported through appropriate maintenance, rather than treated as a one off intervention (Teixeira da Silva et al., 2018; Celletti et al., 2025). This message is also reflected in policy oriented guidance on limiting, mitigating and compensating soil sealing, which frames prevention as the most reliable option where soil functions remain intact (European Commission, 2012). At the same time, the evidence base is still uneven. Many interventions are assessed in a narrow set of contexts, using different indicators and time horizons, which makes it risky to transfer a “best practice” package without local adjustment, monitoring, and, where possible, independent validation (Gómez-Baggethun and Barton, 2013b; Greiner et al., 2017).

6 Discussion and conclusion

This review set out to bring together dispersed evidence on the ecosystem services provided by soils in cities and on the ways in which urbanization reshapes that capacity. Following the research questions stated in the Introduction, we traced links between soil functions and services, summarized the main pressure pathways, and discussed how spatial planning and remote sensing can support urban soil stewardship. Across the literature, one point comes through clearly: in urban areas, soils operate as living infrastructure. When they remain functional, they regulate water, buffer heat, store carbon, and support biodiversity. When they are sealed, compacted or contaminated, those benefits are reduced or lost, often for long periods (Scalenghe and Marsan, 2009; O’Riordan et al., 2021).

Interpreting the evidence in a broader international context, the prominence of regulating services in cities is consistent with earlier syntheses of urban ecosystem services and with the growing policy focus on climate adaptation and flood risk (Gómez-Baggethun and Barton, 2013a; Setälä et al., 2014). Soil sealing stands out as the clearest tipping point, because it interrupts infiltration and biological activity almost entirely. European guidance documents frame sealing as one of the most severe soil threats, and the scientific literature supports that assessment, especially where impervious cover expands into river valleys and other areas that formerly absorbed rainfall (European Commission, 2012; Stolte et al., 2016). In practical terms, the most immediate consequence is hydrological: the loss of permeable surfaces and structure increases runoff peaks and can degrade water quality during intense rainfall (Bolund and Hunhammar, 1999; O’Riordan et al., 2021).

Local environmental conditions and social context then determine how these general relationships play out. Climate sets the background demand. In humid regions, infiltration and retention are often the first concern, while in temperate and continental settings heat mitigation and carbon storage can become equally pressing. Soil type and land use history add another layer; disturbed or imported materials, shallow anthropogenic horizons and chronic compaction may limit rooting depth and water movement even where land is nominally green (Morel et al., 2015; Li et al., 2018). Social factors matter just as much. Maintenance regimes, irrigation practices, vegetation choices, and access to public space influence whether green infrastructure functions as intended and for whom it delivers benefits (Seto et al., 2013; Gómez-Baggethun et al., 2013b). In Central and Eastern Europe, including Poland, the same interplay is visible in urban climate and planning studies, which fits well with the wider European understanding of soil threats and urban heat dynamics (Fortuniak, 2003; European Commission, 2012).

When studies from different settings are read together, a few recurring patterns become clear. In wetter climates, authors tend to focus on infiltration, retention, and controlling peak runoff, especially where sealing expands within catchments (O’Riordan et al., 2021; European Commission, 2012; Stolte et al., 2016). In warmer regions and in dense city centers, the emphasis shifts toward heat exposure and cooling, and the evidence base often draws on surface temperature and vegetation indicators (Fortuniak, 2003; Guha et al., 2018; Zhou et al., 2019). Urban form matters too. Compact inner districts have very limited room for measures that restore permeability, whereas low density growth is closely tied to land take and the steady increase of impervious cover (Seto et al., 2013). What can realistically be monitored and maintained also depends on local resources and institutional capacity, so similar measures can perform quite differently from one city to the next (Gómez-Baggethun and Barton, 2013a; Teixeira da Silva et al., 2018). Unsurprisingly, the literature also points to trade offs that need to be made explicit, for instance between accommodating growth and protecting permeable soils, or between maximizing cooling and taking on long term maintenance commitments (Setälä et al., 2014; European Commission, 2012).

Our synthesis also clarifies both the value and the limits of contemporary monitoring. Remote sensing and Earth observation are now routinely used to map imperviousness, vegetation cover and surface temperature, and these variables can serve as practical proxies for selected services, particularly cooling and runoff regulation (Mulder et al., 2011; Zhou et al., 2019). Indices such as NDVI and NDBI, together with thermal products derived from Landsat or Sentinel data, have been applied across many cities, which increasingly allows comparisons between places using similar workflows (Small and Milesi, 2013; Guha et al., 2018). At the same time, most soil services remain largely below ground. Carbon dynamics, nutrient cycling and biodiversity cannot be inferred reliably from imagery alone and require field measurements for calibration and interpretation (Greiner et al., 2017; Chabrillat et al., 2019). For evaluation and identification tasks in planning, this distinction matters: remote sensing is well suited to indicate where pressures and hotspots are likely, but it cannot by itself confirm functional loss or recovery.

The policy implications follow directly. If soils are treated as infrastructure, spatial planning needs a soil component that goes beyond contamination control. Approaches that integrate soil ecosystem services into planning emphasize prevention of sealing, targeted protection of high value soils, and the use of nature based solutions that keep water in the landscape while providing cooling and biodiversity benefits (Teixeira da Silva et al., 2018; Raymond et al., 2017). European guidance already points in this direction by recommending limitation, mitigation and compensation of sealing. What is often missing at municipal level is an operational bridge between policy aims and everyday decisions on land use and investments (European Commission, 2012; Stolte et al., 2016). Valuation approaches, whether economic or socio-cultural, can support that bridge by translating soil functions into decision relevant terms, but they work best when paired with transparent biophysical indicators (Chee, 2004; Guerry et al., 2015). For transparency, Table 4 summarizes the quantitative statements used across the manuscript and points to the primary sources on which they rely.

Table 4

Table 4. Key quantitative statements used in the manuscript and their primary sources.

Several caveats should be stated clearly. Uncertainty becomes most visible where conclusions rely on indirect indicators and proxies that have not been calibrated against field data. This is still common for belowground processes and biodiversity related outcomes, where remote sensing captures surface signals far more readily than soil properties or soil biota (Mulder et al., 2011; Chabrillat et al., 2019; Finizio et al., 2024). Even for pressures that imagery can map well, such as sealing or vegetation stress, the meaning of a proxy can change with spatial resolution and land cover mixing, so field validation remains important when results are used for assessment and decision making (Greiner et al., 2017; Dean et al., 2024). Another gap is temporal: long term datasets that track recovery after de sealing or structural restoration remain scarce, and reporting is still difficult to compare across climatic zones and socioeconomic settings (O’Riordan et al., 2021; Teixeira da Silva et al., 2018). For these reasons, we do not treat the indicators reported in the literature as universal thresholds. We highlight where evidence is consistent, where it diverges, and where it is still too limited to generalize with confidence. The reviewed studies use different definitions, indicators and spatial scales, so it is not possible to provide a single set of universal thresholds for service loss or to rank cities by soil performance without extensive local standardization (Breure et al., 2012; Greiner et al., 2017). The evidence base is also geographically uneven, and the search was limited to English and Polish publications, which may under-represent findings from other language communities. Finally, publication bias and the scarcity of long term monitoring constrain what can be concluded about recovery trajectories after restoration, particularly for biodiversity and for slow changes in soil structure (O’Riordan et al., 2021).

Taken together, the literature supports a coherent answer to the aims posed in the Introduction. Urban soils provide a broad suite of ecosystem services, with regulating services at the core of urban resilience. Urbanization, particularly sealing, erodes this capacity, but the review also shows that planning choices and management practices can protect and, in some contexts, restore key functions. Remote sensing has matured into a practical tool for identifying priority areas and tracking surface change, yet it needs to be complemented by ground data when decisions depend on below ground processes. Making soil ecosystem services visible in spatial planning and investment decisions is therefore not an optional refinement; it is one of the conditions for sustaining livable cities as urban populations continue to grow (United NationsDepartment of Economic and Social AffairsPopulation Division, 2022; Teixeira da Silva et al., 2018).

Data availability statement

The original contributions presented in the study are included in the article/Supplementary Material, further inquiries can be directed to the corresponding authors.

Author contributions

LP: Writing – original draft, Project administration, Writing – review and editing, Conceptualization, Methodology, Investigation. EG: Writing – original draft, Writing – review and editing, Investigation, Methodology, Visualization, Conceptualization. GS: Funding acquisition, Conceptualization, Writing – review and editing.

Funding

The author(s) declared that financial support was received for this work and/or its publication. Research was performed under NBSOIL (Nature Based Solutions for Soil Management) project, grant number 101091246, financed by Mission Soil of Horizon Europe programme. Research project partly supported by program “Excellence initiative – research university” for the AGH University.

Conflict of interest

The author(s) declared that this work was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Generative AI statement

The author(s) declared that generative AI was not used in the creation of this manuscript.

Any alternative text (alt text) provided alongside figures in this article has been generated by Frontiers with the support of artificial intelligence and reasonable efforts have been made to ensure accuracy, including review by the authors wherever possible. If you identify any issues, please contact us.

Publisher’s note

All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.

Supplementary material

The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fenvs.2025.1734143/full#supplementary-material

References

Adamczyk, A. B., Błażejczyk, K., Baranowski, J., and Kuchcik, M. (2008). “Warunki termiczne aglomeracji warszawskiej,” in Klimat i bioklimat miast. Editors K. Kłysik, J. Wibig, and K. Fortuniak (Łódź, Poland: Wydawnictwo Uniwersytetu Łódzkiego), 11–20.

Google Scholar

Ahmad, N., and Hassan, Q. (2024). Ecosystem services linked to nature-based solutions for resilient and sustainable cities in India. Front. Water 6, 1504492. doi:10.3389/frwa.2024.1504492

CrossRef Full Text | Google Scholar

Ahrné, K., Bengtsson, J., and Elmqvist, T. (2009). Bumble Bees (Bombus spp) along a gradient of increasing urbanization. PLoS One 4 (5), e5574. doi:10.1371/journal.pone.0005574

PubMed Abstract | CrossRef Full Text | Google Scholar

Araya, S. N., Fryjoff-Hung, A., Anderson, A., Viers, J. H., and Ghezzehei, T. A. (2021). Advances in soil moisture retrieval from multispectral remote sensing using UAS and machine learning. Hydrol. Earth. Syst. Sci. 25, 2739–2758. doi:10.5194/hess-25-2739-2021

CrossRef Full Text | Google Scholar

Arkema, K. K., Verutes, G. M., Wood, S. A., Clarke-Samuels, C., Rosado, S., Canto, M., et al. (2015). Embedding ecosystem services in coastal planning leads to better outcomes for people and nature. Proc. Natl. Acad. Sci. U. S. A. 112 (24), 7390–7395. doi:10.1073/pnas.1406483112

PubMed Abstract | CrossRef Full Text | Google Scholar

Baveye, P. C., Baveye, J., and Gowdy, J. (2016). Soil “ecosystem” services and natural capital: critical appraisal of research on uncertain ground. Front. Environ. Sci. 4, 1–49. doi:10.3389/fenvs.2016.00041

CrossRef Full Text | Google Scholar

Blum, W. E. H. (2005). Functions of soil for society and the environment. Rev. Environ. Sci. Biotechnol. 4 (3), 75–79. doi:10.1007/s11157-005-2236-x

CrossRef Full Text | Google Scholar

Bolund, P., and Hunhammar, S. (1999). Ecosystem services in urban areas. Ecol. Econ. 29 (2), 293–301. doi:10.1016/S0921-8009(99)00013-0

CrossRef Full Text | Google Scholar

Boyd, J., and Banzhaf, S. (2007). What are ecosystem services? The need for standardized environmental accounting units. Ecol. Econ. 63 (2-3), 616–626. doi:10.1016/j.ecolecon.2007.01.002

CrossRef Full Text | Google Scholar

Breure, A. M., de Deyn, G. B., Dominati, E., Eglin, T., Hedlund, K., van Orshoven, J., et al. (2012). Ecosystem services: a useful concept for soil policy making. Curr. Opin. Environ. Sustain. 4 (5), 578–585. doi:10.1016/j.cosust.2012.10.010

CrossRef Full Text | Google Scholar

Celletti, S., Poręba, L., Wawer, R., Padoan, E., Comis, S., Bartosiewicz, B., et al. (2025). The potential of nature-based solutions for urban soils: focus on green infrastructure and bioremediation. Front. Environ. Sci. 13, 1634662. doi:10.3389/fenvs.2025.1634662

CrossRef Full Text | Google Scholar

Chabrillat, S., Ben-Dor, E., Cierniewski, J., Gomez, C., Schmid, T., and van Wesemael, B. (2019). Imaging spectroscopy for soil mapping and monitoring. Surv. Geophys. 40 (3), 361–399. doi:10.1007/s10712-019-09524-0

CrossRef Full Text | Google Scholar

Chee, Y. E. (2004). An ecological perspective on valuation of ecosystem services. Biol. Conserv. 120 (4), 549–565. doi:10.1016/j.biocon.2004.03.028

CrossRef Full Text | Google Scholar

Chiesura, A. (2004). The role of urban parks for sustainable city. Landsc. Urban. Plann. 68 (1), 129–138. doi:10.1016/j.landurbplan.2003.08.003

CrossRef Full Text | Google Scholar

Comerford, N. B., Franzluebbers, A. J., Stromberger, M. E., Morris, L., Markewitz, D., and Moore, R. (2013). Assessment and evaluation of soil ecosystem services. Soil Horizons 54 (3), 1–14. doi:10.2136/sh12-10-0028

CrossRef Full Text | Google Scholar

Constanza, R., de Groot, R., Sutton, P., van der Ploeg, S., Anderson, S. J., Kubiszewski, I., et al. (2014). Changes in the global value of ecosystem services. Global. Environ. Change 26 (1), 152–158. doi:10.1016/j.gloenvcha.2014.04.002

CrossRef Full Text | Google Scholar

Costanza, R., de Groot, R., Braat, L., Kubiszewski, I., Fioramonti, L., Sutton, P., et al. (2017). Twenty years of ecosystem services: how far have we come and how far do we still need to go? Ecosyst. Serv. 28, 1–16. doi:10.1016/j.ecoser.2017.09.008

CrossRef Full Text | Google Scholar

Czarnecki, A., and Lewnadowska-Czarnecka, A. (2007). Zastosowanie koncepcji ekosystemowych do zarządzania przestrzenią. Czas. Techn. 7-A, 65–71. (In Polish).

Google Scholar

de Groot, R. S., Wilson, M. A., and Boumans, R. M. J. (2002). A typology for the classification, description and valuation of ecosystem functions, goods and services. Ecol. Econ. 41 (3), 393–408. doi:10.1016/S0921-8009(02)00089-7

CrossRef Full Text | Google Scholar

Dean, J. R., Ahmed, S., Cheung, W., Salaudeen, I., Reynolds, M., Bowerbank, S. L., et al. (2024). Use of remote sensing to assess vegetative stress as a proxy for soil contamination. Environ. Sci. Process. & Impacts 26, 161–176. doi:10.1039/D3EM00480E

PubMed Abstract | CrossRef Full Text | Google Scholar

Dominati, E., Patterson, M., and Mackay, A. (2010). A framework for classifying and quantifying the natural capital and ecosystem services of soils. Ecol. Econ. 69 (9), 1858–1868. doi:10.1016/j.ecolecon.2010.05.002

CrossRef Full Text | Google Scholar

Elmqvist, T., Redman, C. L., Barthel, S., and Constanza, R. (2013). “History of urbanization and the missing ecology,” in Urbanization, biodiversity and ecosystem services: challenges and opportunities: a global assessment. Editors T. Elmqvist, M. Fragkias, J. Goodness, B. Güneralp, P. J. Marcotullio, and R. I. McDonald (Dordrecht: Springer).

CrossRef Full Text | Google Scholar

European Commission (2012). “Wytyczne dotyczące najlepszych praktyk w zakresie ograniczania,” in Łagodzenia i kompensowania procesu zasklepiania gleb. Luxembourg: Publications Office of the European Union. (In Polish and English).

Google Scholar

European Commission (2025). Green infrastructure — enhancing Europe’s natural capital.

Google Scholar

Famielec, J., Górka, K., Stuczyński, T., and Wołkowicz, S. (2007). Oszacowanie kosztów wynikających z wdrażania w Polsce wymagań zawartych w projekcie dyrektywy Parlamentu Europejskiego i rady ustanawiającej ramy dla ochrony gleb oraz zmieniającej dyrektywę 2004/35/WE. Report to Polish Ministry of Environment (Warsaw) (In Polish)

Google Scholar

Fang, C. F., and Ling, D.-L. (2003). Investigation of the noise reduction provided by tree belts. Landsc. Urban. Plann. 63 (4), 187–195. doi:10.1016/S0169-2046(02)00190-1

CrossRef Full Text | Google Scholar

Finizio, M., Pontieri, F., Bottaro, C., Di Febbraro, M., Innangi, M., Sona, G., et al. (2024). Remote sensing for urban biodiversity: a review and meta-analysis. Remote Sens. 16 (23), 4483. doi:10.3390/rs16234483

CrossRef Full Text | Google Scholar

Food and Agriculture Organization of the United Nations (FAO) (2015). Revised world soil charter. (FAO: rome, Italy) food and agriculture organization of the united nations and intergovernmental technical panel on soils (2020). State of knowledge of soil biodiversity: status, challenges and potentialities. Rome, Italy: FAO. doi:10.4060/cb1928en

CrossRef Full Text | Google Scholar

Fortuniak, K. (2003). Miejska wyspa ciepła. Podstawy energetyczne, studia eksperymentalne, modele numeryczne i statystyczne. Łódź, Poland: Wydawnictwo Uniwersytetu Łódzkiego.

Google Scholar

Gerst, F., Bubenzer, O., and Mächtle, B. (2011). “Climatic impacts of urban soil consumption – review and system model,”Rep. EU-project Urban-SMS. WP6 Awarness & Acceptance. Stuttgart: Ruprecht-Karls-Universität Heidelberg.

Google Scholar

Głowienka, E., Malinverni, E. S., Kucza, M., Michałowska, K., and Sanità, M. (2025a). Satellite and ground-based data to monitor urban heat Islands. Cases Study Pol. Italian Cities. ISPRS Archives. XLVIII-G-2025, 521–527. doi:10.5194/isprs-archives-XLVIII-G-2025-521-2025

CrossRef Full Text | Google Scholar

Głowienka, E., Malinverni, E. S., Sanità, M., Michałowska, K., and Kucza, M. (2025b). Harmonizing satellite thermal data with ground-based observations for climate long-term monitoring. ISPRS Arch. XLVIII-M-7–2025, 127–132. doi:10.5194/isprs-archives-XLVIII-M-7-2025-127-2025

CrossRef Full Text | Google Scholar

Goličnik Marušić, B., Dremel, M., and Ravnikar, Ž. (2023). A frame of understanding to better link nature-based solutions and urban planning. Environ. Sci. Policy 146, 47–56. doi:10.1016/j.envsci.2023.05.005

CrossRef Full Text | Google Scholar

Gómez-Baggethun, E., and Barton, D. N. (2013a). Classifying and valuing ecosystem services for urban planning. Ecol. Econ. 86, 235–245. doi:10.1016/j.ecolecon.2012.08.019

CrossRef Full Text | Google Scholar

Gómez-Baggethun, E., Gren, Å., Barton, D. N., Langemeyer, J., McPhearson, T., O’Farrell, P., et al. (2013b). Urban ecosystem services,” in urbanization, biodiversity and ecosystem services: challenges and opportunities: a global assessment. Editors T. Elmqvist, M. Fragkias, J. Goodness, B. Güneralp, P. J. Marcotullio, R. I. McDonaldet al. (Springer Dordrecht).

Google Scholar

Gregory, A. S., Ritz, K., McGrath, S. P., Quinton, J. N., Goulding, K. W. T., Jones, R. J. A., et al. (2015). A review of the impacts of degradation threats on soil properties in the UK. Soil. Use. Manage. 31 (S1), 1–15. doi:10.1111/sum.12118

PubMed Abstract | CrossRef Full Text | Google Scholar

Greiner, L., Keller, A., Grêt-Regamey, A., and Papritz, A. (2017). Soil function assessment: review of methods for quantifying the contributions of soils to ecosystem services. Land Use Policy 69, 224–237. doi:10.1016/j.landusepol.2017.06.025

CrossRef Full Text | Google Scholar

Gruszczyński, S. (2014). Zmiany w środowisku glebowym i ich skutki. Przyszłość Świat-Europa-Polska (2), 36–63. (In Polish).

Google Scholar

Guerry, A. D., Polasky, S., Lubchenco, J., Chaplin-Kramer, R., Daily, G. C., Griffin, R., et al. (2015). Natural capital and ecosystem services informing decisions: from promise to practice. Proc. Natl. Acad. Sci. U. S. A. 112 (24), 7348–7355. doi:10.1073/pnas.1503751112

PubMed Abstract | CrossRef Full Text | Google Scholar

Guha, S., Govil, H., Dey, A., and Gill, N. (2018). Analytical study of land surface temperature with NDVI and NDBI using landsat 8 OLI and TIRS data in florence and Naples city, Italy. Eur. J. Remote. Sens. 51, 667–678. doi:10.1080/22797254.2018.1474494

CrossRef Full Text | Google Scholar

Guilland, C., Maron, P. A., Damas, O., and Ranjard, L. (2018). Biodiversity of urban soils for sustainable cities. Environ. Chem. Lett. 16 (4), 1267–1282. doi:10.1007/s10311-018-0751-6

CrossRef Full Text | Google Scholar

Guo, Y., Zhang, H., Li, Q., Lin, Y., and Michalski, J. (2022). New morphological features for urban tree species identification using LiDAR point clouds. Urban For Urban Green 71, 127558. doi:10.1016/j.ufug.2022.127558

CrossRef Full Text | Google Scholar

Hejmanowska, B., Głowienka, E., and Michałowska, K. (2016). “Free satellite imagery for monitoring reclaimed Sulphur mining region tarnobrzeg, Poland,” in 2016 Baltic geodetic congress (Geomatics) (IEEE), 134–139. doi:10.1109/BGC.Geomatics.2016.32

CrossRef Full Text | Google Scholar

Iniesta-Arandia, I., García-Llorente, M., Aguilera, P. A., Montes, C., and Martín-López, B. (2014). Socio-cultural valuation of ecosystem services: uncovering the inks between values, drivers of change, and human well-being. Ecol. Econ. 108, 36–48. doi:10.1016/j.ecolecon.2014.09.028

CrossRef Full Text | Google Scholar

Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (2019). in Summary for policymakers of the global assessment report on biodiversity and ecosystem services. Editors E. S. Brondizio, J. Settele, S. Díaz, and H. T. Ngo (Bonn, Germany: IPBES Secretariat). doi:10.5281/zenodo.3553579

CrossRef Full Text | Google Scholar

Janhäll, S. (2015). Review on urban vegetation and particle air pollution – deposition and dispersion. Atmos. Environ. 105, 130–137. doi:10.1016/j.atmosenv.2015.01.052

CrossRef Full Text | Google Scholar

Jha, P., Joy, M. S., Yadav, P. K., Begam, S., and Bansal, T. (2024). Detecting the role of urban green parks in thermal comfort and public health for sustainable urban planning in Delhi. Discov. Public Health 21, 236. doi:10.1186/s12982-024-00368-7

CrossRef Full Text | Google Scholar

Kamiński, J., Głowienka, E., Soszyński, D., Trzaskowska, E., Stuczyński, T., Siebielec, G., et al. (2025). Integrating expert assessments and spectral methods to evaluate visual attractiveness and ecosystem services of urban informal green spaces in the context of climate adaptation. Sustainability 17, 1349. doi:10.3390/su17041349

CrossRef Full Text | Google Scholar

Kara, Y., Yavuz, V., and Lupo, A. R. (2025). Multi-index assessment of surface Urban heat Island (SUHI) dynamics in Samsun using google Earth engine. Atmosphere 16, 712. doi:10.3390/atmos16060712

CrossRef Full Text | Google Scholar

Keesstra, S. D., Geissen, V., Mosse, K., Piirainen, S., Scudiero, E., Leistra, M., et al. (2012). Soil as a filter for groundwater quality. Curr. Opin. Environ. Sustain. 4 (5), 507–516. doi:10.1016/j.cosust.2012.10.007

CrossRef Full Text | Google Scholar

Kong, Y., Jimenez, K., Lee, C. M., Hokanson, K., Maruya, K., Sutula, M., et al. (2025). Monitoring coastal water turbidity using Sentinel-2: a case study in Los Angeles. Remote Sens. 17 (2), 201. doi:10.3390/rs17020201

CrossRef Full Text | Google Scholar

Li, G., Sun, G.-X., Ren, Y., Luo, X.-S., and Zhu, Y.-G. (2018). Urban soil and human health: a review. Eur. J. Soil. Sci. 69 (1), 196–215. doi:10.1111/ejss.12518

CrossRef Full Text | Google Scholar

Lovynska, V., Bayat, B., Bol, R., Šimůnek, J., Kuzyakov, Y., Mason, K. E., et al. (2024). Monitoring heavy metals and metalloids in soils and vegetation by remote sensing: a review. Remote Sens. 16 (17), 3221. doi:10.3390/rs16173221

CrossRef Full Text | Google Scholar

Łuczyszyn, A., and Łuczyszyn, M. (2018). Urbanizacja jako element rozwoju miast. Pr. Nauk. Uniw. Ekon. We Wrocławiu. 502, 53–61. (In Polish). doi:10.15611/pn.2018.502.05

CrossRef Full Text | Google Scholar

Mangafić, A., Oštir, K., Kolar, M., and Zupan, M. (2025). Hyperspectral soil heavy metal prediction via privileged-informed residual correction. Remote Sens. 17 (12), 1987. doi:10.3390/rs17121987

CrossRef Full Text | Google Scholar

McDonald, R. I., Marcotullio, P. J., and Güneralp, B. (2013). “Urbanization and global trends in biodiversity and ecosystem services,” in Urbanization, biodiversity and ecosystem services: challenges and opportunities: a global assessment. Editors T. Elmqvist, M. Fragkias, J. Goodness, B. Güneralp, P. J. Marcotullio, and R. I. McDonald (Dordrecht: Springer).

CrossRef Full Text | Google Scholar

Meulen, S., Maring, L., Bartkowski, B., Hagemann, N., Arrüe, J., Playán, E., et al. (2018). Mapping and assessment of ecosystems and their services: soil ecosystems.

Google Scholar

Michałowska, K., Głowienka, E., and Hejmanowska, B. (2016). Temporal satellite images in the process of automatic efficient detection of changes of the Baltic sea coastal zone. IOP Conf. Ser. Earth Environ. Sci. 44, 042019. doi:10.1088/1755-1315/44/4/042019

CrossRef Full Text | Google Scholar

Millennium Ecosystem Assessment (2005). Ecosystems and human well-being: synthesis. Washington, DC, United States of America: Island Press.

Google Scholar

Morel, J. L., Chenu, C., and Lorenz, K. (2015). Ecosystem services provided by soils of urban, industrial, traffic, mining and military areas (SUITMAs). J. Soils. Sediments. 15 (8), 1659–1666. doi:10.1007/s11368-014-0926-0

CrossRef Full Text | Google Scholar

Mulder, V. L., de Bruin, S., Schaepman, M. E., and Mayr, T. (2011). The use of remote sensing in soil and terrain mapping – a review. Geoderma 162 (1–2), 1–19. doi:10.1016/j.geoderma.2010.12.018

CrossRef Full Text | Google Scholar

Nawaz, M. F., Bourrié, G., and Trolard, F. (2013). Soil compaction impact and modelling: a review. Agron. Sustain. Dev. 33 (2), 291–309. doi:10.1007/s13593-011-0071-8

CrossRef Full Text | Google Scholar

Nidzgorska-Lencewicz, J., and Mąkosza, A. (2016). Specyficzne cechy klimatu miasta w aspekcie zdrowia człowieka. Kosm. Probl. Nauk. Biol. 65 (4), 637–645. (In Polish).

Google Scholar

Orlova, K. S., and Savin, I. Y. (2024). Ecosystem services provided by urban soils and their assessment: a review. Eurasian. Soil. Sci. 57 (9), 1072–1083. doi:10.31857/S0032180X24060083

CrossRef Full Text | Google Scholar

Osińska-Skotak, K. (2002). Przydatność zdjęć termalnych dla planowania przestrzennego w obszarach miejskich. Arch. Fotogram. Kartogr. I Teledetekcji 12B, 307–324. (In Polish).

Google Scholar

O’Riordan, R., Davies, J., Stevens, C., Quinton, J., and Boyko, C. (2021). The ecosystem services of urban soils: a review. Geoderma 395, 115076. doi:10.1016/j.geoderma.2021.115076

CrossRef Full Text | Google Scholar

Pouyat, R. V., Yesilonis, I. D., and Nowak, D. J. (2006). Carbon storage by urban soils in the United States. J. Enviro. Qual. 35 (4), 1566–1575. doi:10.2134/jeq2005.0215

PubMed Abstract | CrossRef Full Text | Google Scholar

Randin, C. F., Ashcroft, M. B., Bolliger, J., Cavender-Bares, J., Coops, N. C., Drever, C. R., et al. (2020). Monitoring biodiversity in the Anthropocene using remote sensing in species distribution models. Remote. Sens. Environ. 239, 111626. doi:10.1016/j.rse.2019.111626

CrossRef Full Text | Google Scholar

Raymond, C. M., Frantzeskaki, N., Kabisch, N., Berry, P., Breil, M., Nita, M. R., et al. (2017). A framework for assessing and implementing the co-benefits of nature-based solutions in urban areas. Environ. Sci. Policy 77, 15–24. doi:10.1016/j.envsci.2017.07.008

CrossRef Full Text | Google Scholar

Rocchini, D., Boyd, D. S., Féret, J.-B., Foody, G. M., He, K. S., Lausch, A., et al. (2018). Satellite remote sensing to monitor species diversity: potential and pitfalls. Remote. Sens. Ecol. Conserv. 4 (1), 7–20. doi:10.1002/rse2.53

CrossRef Full Text | Google Scholar

Scalenghe, R., and Marsan, F. A. (2009). The anthropogenic sealing of soils in urban areas. Landsc. Urban Plann 90, 1–10. doi:10.1016/j.landurbplan.2008.10.011

CrossRef Full Text | Google Scholar

Setälä, H., Bardgett, R. D., Birkhofer, K., Brady, M., Byrne, L., de Ruiter, P. C., et al. (2014). Urban and agricultural soils: conflicts and trade-offs in the optimization of ecosystem services. Urban Ecosyst. 17 (1), 239–253. doi:10.1007/s11252-013-0311-6

CrossRef Full Text | Google Scholar

Seto, K. C., Parnell, S., and Elmqvist, T. (2013). “A global outlook on urbanization,” in Urbanization, biodiversity and ecosystem services: challenges and opportunities: a global assessment. Editors T. Elmqvist, M. Fragkias, J. Goodness, B. Güneralp, P. J. Marcotullio, and R. I. McDonald (Dordrecht: Springer).

CrossRef Full Text | Google Scholar

Shao, Z., Cheng, T., Fu, H., Li, D., and Huang, X. (2023). Emerging issues in mapping urban impervious surfaces using high-resolution remote sensing images. Remote Sens. 15 (10), 2562. doi:10.3390/rs15102562

CrossRef Full Text | Google Scholar

Sikora, S. (2008). “Stres gorąca jako istotne ograniczenie komfortu życia w mieście: Jego natężenie i częstość pojawiania się na przykładzie Wrocławia,” in Klimat i bioklimat miast. Editors K. Kłysik, J. Wibig, and K. Fortuniak (Łódź, Poland: Wydawnictwo Uniwersytetu Łódzkiego), 343–352.

Google Scholar

Small, C., and Milesi, C. (2013). Multi-scale standardized spectra mixture models. Remote Sens. Environ. 136, 442–454. doi:10.1016/j.rse.2013.05.024

CrossRef Full Text | Google Scholar

Smreczak, B., Ukalska-Jaruga, A., Łysiak, M., Strzelecka, J., Niedźwiecki, J., and Sobich, D. (2017). Funkcje, jakość i usługi ekosystemowe gleb. SiR 54 (8), 9–23. (In Polish).

Google Scholar

Solecki, W., and Marcotullio, P. J. (2013). “Climate change and urban biodiversity vulnerability,” in Urbanization, biodiversity and ecosystem services: challenges and opportunities: a global assessment. Editors T. Elmqvist, M. Fragkias, J. Goodness, B. Güneralp, P. J. Marcotullio, and R. I. McDonald (Dordrecht: Springer).

CrossRef Full Text | Google Scholar

Solon, J. (2008). Koncepcja “Ecosystem Services” i jej zastosowanie w badaniach ekologiczno-krajobrazowych. Probl. Ekol. Kraj. 21, 25–44. (In Polish).

Google Scholar

Soltani, A., and Sharifi, E. (2017). Daily variation of urban heat island effect and its correlations to urban greenery: a case study of Adelaide. Front. Archit. Res. 6 (4), 529–538. doi:10.1016/j.foar.2017.08.001

CrossRef Full Text | Google Scholar

J. Stolte, M. Tesfai, L. Øygarden, S. Kværnø, J. Keizer, and F. Verheijen (2016). Soil threats in Europe (Brussels: EUR 27607 EN, European Commission). doi:10.2788/828742

CrossRef Full Text | Google Scholar

Sudra, P. (2015). Usługi ekosystemowe na tle wybranych koncepcji ekologii miasta. Człow. I Środowisko 39, 61–73. (In Polish).

Google Scholar

Tan, M., Zhang, X., Luo, W., and Hao, M. (2023). Deep learning based spatial distribution estimation of soil Pb using multi-phase multispectral remote sensing images in a mining area. Land 12, 1789. doi:10.3390/land12091789

CrossRef Full Text | Google Scholar

Teixeira da Silva, R., Fleskens, L., van Delden, H., and van der Ploeg, M. (2018). Incorporating soil ecosystem services into urban planning: status, challenges and opportunities. Landsc. Ecol. 33 (7), 1087–1102. doi:10.1007/s10980-018-0652-x

CrossRef Full Text | Google Scholar

Polskie Towarzystwo Gleboznawcze, (2019). Systematyka gleb Polski (soil Classification of Poland). Warsaw, Pol. Pol. Soc. Soil Sci.

Google Scholar

L. Trząski (2015). Przestrzenny i środowiskowy wymiar zrównoważonego rozwoju terenów zurbanizowanych. Katowice, Poland: główny Instytut Górnictwa.

Google Scholar

United Nations, Department of Economic and Social Affairs, Population Division (2022). World urbanisation prospects: the 2022 revision. New York, United States of America: United Nations.

Google Scholar

F. Woch (2015). Wademekum klasyfikatora gleb. Pulawy, Poland: IUNG-PIB.

Google Scholar

S. Zawadzki (1999). Gleboznawstwo. Warsaw, Poland: państwowe wydawnictwo rolnicze i leśne.

Google Scholar

Zhao, L., Fan, X., and Hong, T. (2025). Urban heat island effect: remote sensing monitoring and assessment—Methods, applications, and future directions. Atmosphere 16 (7), 791. doi:10.3390/atmos16070791

CrossRef Full Text | Google Scholar

Zhou, D., Xiao, J., Bonafoni, S., Zhang, L., Wei, J., Zhang, G., et al. (2019). Satellite remote sensing of surface urban heat islands: progress, challenges, and perspectives. Remote Sens. 11 (1), 48. doi:10.3390/rs11010048

CrossRef Full Text | Google Scholar