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.

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.

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.

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.

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.

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.

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.

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.

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).

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.

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.

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

regulating services,

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.

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,

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.

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).

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.

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).

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).

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).

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).

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).