Soil–Plant Indicators for Assessing Nutrient Cycling and Ecosystem Functionality in Urban Forestry

Abstract

Nature-based solutions (NbS) are multidimensional, resource-efficient, and sustainable growth approaches to cope with current challenges, including biodiversity and carbon loss, pollution, climate change and land degradation. Amongst NbS, urban forestry is an important tool to enhance environmental resilience and sustainability, providing useful ecosystem services for human well-being. In this context, using suitable soil and plant indicators allows us to evaluate the efficiency of urban forestry in sustaining ecosystem functionality. Effective indicators should be sensitive to environmental changes and representative of ecological processes. Many studies focus on the selection of soil or plant indicators. The prior investigations considered soil–plant interaction and the related complex heterarchical and bidirectional effects involving plant strategy and soil biota. The choice and the use of indicators related to the soil–plant system could be an innovative strategy to better assess the following: (1) the ability of soil to support healthy plants and their ability to improve air quality; (2) the effect of urban forestry on ecological processes, in particular carbon and nutrient cycles. This review investigates the suitability of soil–plant system indicators related to nutrient cycles, e.g., ecological stoichiometry, enzyme activity and stoichiometry, and carbon and nitrogen stable isotopes, as valuable tools for planning and evaluating the effectiveness of urban forestry interventions.

1. Introduction

1.1. Urban Environment Degradation

The remarkable growth of urban areas since the conclusion of World War II has seemed unrelenting. The relative growth rate is double that of the global population, attributed to a rising percentage of the world’s populace residing in urban areas [1] and to lifestyle alterations associated with the average increase in wealth generation [2,3]. The adverse effects of urban expansion on biotopes, climate change, flood risk, and agricultural output, among other factors, have been extensively studied in numerous research articles globally over the past two decades [4]. Urban development negatively affects ecosystem functions and services, e.g., soil degradation, pollution, carbon and nutrient cycle alteration, and atmospheric deposition [5,6].

The significance of soil in urban environments has also been the subject of heightened research in recent years, as its functions support numerous ecosystem services (ESs) in urban areas, such as microclimate regulation, water management, biodiversity support, and carbon sequestration [7,8,9,10]. However, soil’s ability to provide ES strictly depends on the soil conditions and health [11,12]. Human activities such as construction, industrial pollution, fossil fuel combustion, waste inputs, and fertilisation lead to soil’s physical, chemical, and biological degradation, impeding the soil’s ability to provide ESs [13]. The main physical alteration includes the destruction of soil structure and loss of total and air porosity, lowering of infiltration rate, and saturated conductivity. The decline in adequate water amount and water-adjusting ability, as well as the rise in soil strength, limits soil biota activity and nutrient cycles, and leads to a higher resistance to the root expansion of urban plants [14,15,16,17]. Concerning the chemical alterations, a slightly extremely strong alkaline pH generally characterises urban soil due to building wastes (concrete and cement) that release calcareous solution [14]; changes in pH can affect the activity of soil decomposers, which can have an impact on soil structure formation and on nutrient cycles [5,18,19,20].

Several studies have demonstrated that the concentration of heavy metals in urban soil is higher than in nearby soils, and it varies based on industry types. However, most research shows that the primary polluting elements are Cu, Zn, Pb, and Hg, known as “urban heavy metals”. However, heavy metals are just one facet of contamination of the urban soil, which also represents a sink of organic contaminants such as polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), and polychlorinated naphthalenes (PCNs) produced by human activity [14]. Urbanisation also results in significant changes in soil carbon stocks, altering the biochemical process and the dynamic of soil organic matter (SOM) [21,22,23], impacting the global climate [24,25,26,27,28,29,30,31].

Urban areas are also major source of atmospheric carbon mono- and dioxide (CO and CO₂), nitrogen oxides (NOx), sulphur dioxide (SO₂), particulate matter (PM), and volatile organic compounds (VOCs), mainly from transport, manufacturing, and power-generating industries, contributing up to 70% of total GHG emissions [32,33,34,35,36]. The impact of urbanisation on air pollution is a growing environmental concern, with wide-ranging impacts on human health and ecosystems. In addition to having harmful effects on human health (e.g., cardiovascular and respiratory diseases), air pollution also affects soil nutrient cycles and biodiversity [37,38], which are essential to maintain healthy ecosystems and support plant life [38,39,40].

1.2. Nature-Based Solutions

Nature-based Solutions (NbS) are essential for alleviating the detrimental impacts of urbanisation on air and soil health. The International Union for Conservation of Nature (IUCN) has defined NbS as “actions to protect, sustainably manage and restore (create) natural or modified ecosystems that address societal challenges (including urban ones) effectively and adaptively, simultaneously providing human well-being and biodiversity benefits” [41]. Successively, the European Commission defined NbS as “Solutions that are inspired and supported by nature, which are cost-effective, simultaneously provide environmental, social, and economic benefits and help build resilience” [42]. In both definitions, NbS are recognised for their multidimensional resource-efficient and sustainable growth approach, capable of addressing local challenges such as biodiversity and carbon loss, pollution, climate change, desertification, and land degradation, and their interlinkages with human well-being, including health and inclusivity [43,44]. According to the IPCC report, NbS are one of the best methods for lowering carbon emissions and are considered one of the key principles of society’s sustainable transition [45]. They have been highlighted as an important concept in several policy publications, including the Kunming-Montreal Global Biodiversity Framework, aimed at transforming society’s relationship with biodiversity by 2030 and ensuring that the shared vision of living in harmony with nature is achieved by 2050 [46]. Furthermore, the Ramsar Convention on Wetlands COP14 in 2022 recognised that NbS can significantly contribute to climate action by improving adaptation and resilience to climate extremes and mitigating climate change and its impacts. It also appears very clear that NbS are essential to achieve the ambitious targets set by the European Nature Restoration Law [47,48] and that EU Member States are required to develop tailored national restoration plans that contribute to EU-wide nature recovery. The Nature Restoration Law represents the first continent-wide step forward in supporting the restoration of terrestrial, marine, and urban ecosystems into the broader context of the EU Biodiversity Strategy for 2030 [49]. It establishes legal requirements for the large-scale restoration of various habitats to restore all ecosystems by 2050 and at least 20% of the EU’s land and sea by 2030. The achievement of the 1.5° Celsius global warming target and biodiversity are dependent upon its implementation.

Regarding urban areas, this law aims to stop the net loss of green spaces and tree cover by 2030, in tandem with the European Green Deal. Furthermore, the law recommends expanding the overall national urban green cover by 3% by 2040 and by 5% by 2050 concerning 2021 levels, focusing on the net gain of urban green space integrated into existing and new building stock and infrastructure development. A minimum of 10% tree canopy cover in every EU city, town, and suburb by 2050 is another urban green target of the law. After 2030, Member States must establish a “satisfactory level” of increasing urban green cover, assessed every six years [50].

The design and application of NbS in the urban context as efficient engineered green measures allows for improved urban resilience and sustainability and provides numerous ESs, such as flood prevention, microclimate and air quality regulation, and enhanced biodiversity [51,52]. Research has shown that implementing green infrastructure on a large scale can effectively respond to climate change [53]. Through locally appropriate, resource-efficient, and systemic interventions, NbS can increase the diversity of nature and its features and processes in cities and landscapes, offering the opportunity to reconcile nature and for people to coexist in urban areas.

In this regard, urban forestry can be considered as an important NbS tool to provide useful ecosystem services for human well-being. Urban forestry, defined as the management and cultivation of tree populations within urban settings, is an essential NbS for improving the environmental sustainability of cities. Its broadest definition encompasses all urban green space management, including parks, squares, gardens, and forests. It combines arboriculture, ornamental horticulture, and forest management. Urban forestry can mitigate some of the most pressing environmental and social challenges that cities face, including climate regulation, air quality improvement, stormwater management, and providing recreational spaces for urban populations [54,55]. It is pivotal in regulating soil ecosystem processes and functions, such as microbial activity, plant–microbe interactions, and organic carbon sequestration [56].

In some cases, altered urban soil conditions (e.g., limited organic carbon and high compaction) can affect the feasibility of urban forestry. The use of recycling material, for example, in tree cultivation and planting, can contribute to alleviating plant stress in transplanting [57,58]. Healthy soils can promote the nutrition, growth, and tolerance of urban plants to both biotic and abiotic stress, which are essential conditions for ensuring the success of forestry interventions in an urban context [59]. On the other hand, healthy plants support physical, chemical and biological soil fertility, providing high-quality litter and root exudates.

The forest ecological equilibrium depends on soil microorganisms [60]. They establish a complex web of interaction with plant roots, impacting soil structure, nutrient cycling, and overall soil–plant system health [61]. Additionally, microbes aid in the breakdown of organic debris, returning vital nutrients to the soil and making it easier for plants to absorb them [62]. The ES provided by trees increases with plant size and age. Moreover, increasing plant biodiversity is a key strategy for improving the effectiveness of land restoration in urban areas, since diverse plant communities can better foster above- and belowground interactions [63,64,65]. It has been demonstrated that the composition of soil microbial and animal communities and the rate of microbial litter decomposition are positively affected by plant biodiversity [66]. In addition, a more biodiverse soil microbial population provides a range of decomposition functions that can improve the stability of the soil ecosystem, the amount of available carbon substrates, and the resistance and resilience to disturbance [67].

Thus, implementing strategies for maintaining healthy soil–plant systems can contribute to urban resilience and promote the development of trees and their ecosystem services, while mitigating potential disservices such as pest management, vegetal material falls, and infrastructure damages [68]. Understanding the interactions between plants and soil health in urban forestry will help us underscore how forest ecosystems function and maintain themselves.

1.3. Soil–Plant Health Indicators

Given the complexity of NbS processes, indicator valuations are continually being developed and optimised. However, poor investigations are carried out on selected indicators that are representative of ecosystem processes. According to several authors [69,70], soil health indicators should be sensitive to land use and management changes while also reflecting soil multifunctionality, which is the soil’s capacity to provide multiple functions or services simultaneously.

Soil health concerns biological organisms, their functions, and their activities in soil [69]. This link implies that microbial diversity and activity are of greater importance in better understanding drivers and interactions and tracking the progress of NbS interventions. In particular, soil enzymes and stoichiometry estimate nutrient availability in the environment, as microbial limitations should reflect plant nutrient limitation [71], resulting in promising indicators of the soil–plant system. Although soil biological indicators are more sensitive in describing soil health, function, and degradation (because they react quickly to environmental perturbations), chemical and physical indicators are the most often used [72]. In this context, soil organic matter (SOM) is an important, widely studied indicator of soil health because it serves multiple functions and affects, through breakdown and transformation processes, other physical, chemical, and biological properties of the soil ecosystems [73,74,75]. Further insights into the origin and dynamics of SOM and nutrient cycles can be obtained by evaluating the carbon and nitrogen isotopic composition of organic matter (i.e., δ¹³C and δ¹⁵N), as well as the elemental stoichiometry in plant, litter, and soil matrices.

Using soil–plant indicators could allow us to understand better the role of soil in supporting healthy plants that provide ESs and urban forestry effects in ecological processes (e.g., carbon and nutrient cycles).

To this end, in this review, we highlight the potential of some indicators of the soil–plant system that are representative of carbon and nutrient cycles (ecological stoichiometry, enzyme activity and stoichiometry, carbon and nitrogen stable isotopes) to assess urban forestry efficiency from a complex ecosystem perspective. The selection, validation, and measurement of these standardised soil–plant indicators will enable the effective planning and maintenance of long-term urban forestry (Figure 1).

2. Materials and Methods

A systematic search of case studies in urban areas was performed, where soil–plant indicators linked to ecological stoichiometry, enzyme activity and stoichiometry, and carbon and nitrogen stable isotopes were investigated. The database Scopus was searched and the search criteria with relative keywords reported in Figure 2 were applied.

3. Suitable Indicators of Soil–Plant System Health

3.1. Ecological Stoichiometry

The ecological stoichiometry approach links the elemental composition, such as the C:N:P of water, soil, and organisms, with ecological processes and traits, e.g., allowing the prediction of nutrient cycles in different environments [76,77]. Table S1 and Figure 3 show the main case studies on ecological stoichiometry and their worldwide distribution. The element stoichiometry in soil can be a suitable indicator of carbon and nutrient dynamics and limitations in NbS design, application, and monitoring. This approach contributes to the selection of the best management practises for creating a healthier urban forest, enabling the related ecosystem services to be assessed.

Microorganisms are the main drivers of carbon and nutrient cycles, primarily through organic matter degradation [78]. The C:N:P ratios in microbial biomass reflect the energy and nutrient needs for microbial growth [77,78]. The elemental stoichiometry of soil resources (e.g., plant litter) used by microorganisms varies amongst habitats and environmental conditions; thus, differences in the C:N:P ratios between microbial biomass and soil resource (mismatches) may occur, resulting in microorganisms not finding the sufficient amount of energy or nutrients they need for their metabolism. In particular, the leaf stoichiometry variability mainly relates to differences amongst species in life forms, nutrient allocation strategy, and photosynthetic performance [77,79]. These mismatches in C:N:P between soil resources and microorganisms can affect the microbial processes (mineralization and immobilisation) through feedback on nutrient availabilities [77]. Specifically, resources with high carbon and nutrient ratios can be related to microbial nutrient limitation, leading to low nutrient mineralisation, high nutrient immobilisation and respiration, and reduced microbial carbon sequestration potential. The low mineralisation process can negatively affect the nutrient availability of plants [71]. On the contrary, the resources’ low carbon and nutrient ratios can increase nutrient mineralisation and reduce microbial respiration, reducing microbial carbon sequestration. The high mineralisation process can lead to nutrient losses [77]. For example, as the carbon and nitrogen cycles tend to be intercorrelated, a balance between plant C input and microbial C output leads to an efficient nitrogen cycle and soil carbon sequestration [80]. The soil stoichiometry can indicate nutrient limitations for plants and microorganisms, as seen in a natural vegetation restoration study by Xue et al. [81]. Specifically, the imbalance of C:N can influence microbial needs, affecting plant nutrition over time.

Plant litter and soil stoichiometry alteration is expected in urban areas compared to natural areas. In particular, the enhanced atmospheric nitrogen deposition, the soil’s physical disturbance, and the limited organic input lower the organic carbon and C:N ratio [82,83]. The atmospheric nitrogen depositions induce nutrient unbalance, e.g., high N:P in plants, litter, and soil, as well as the presence of anthropogenic waste deposition (e.g., construction residues), which can higher the soil phosphorus concentration, lowering C:P and N:P [84,85]. However, Boccuzzi et al. [86] noticed a progressive N enrichment and P reduction over time in the urban ecosystems. The altered soil condition (e.g., low organic matter content, low C:N) can increase the nitrogen cycle and organic matter degradation, leading to a possible loss of soil nutrients [87]. Ecological stoichiometry can be a promising tool for predicting the effect of NbS application on nutrient cycles in urban areas. The soil C:N:P stoichiometry can be a suitable indicator in assessing the impact of tree species on soil conditions [88]. The urban forest types can distinctly affect the soil C:N:P due to the differences in nutrient uptake and organic inputs (litter and root exudates) [88]. In addition, the time elapsed since planting can increase the content of recalcitrant compounds and reduce the soil’s labile organic carbon; thus, C:N and C:P result in sensitive indicators for labile organic carbon dynamics [89]. The tree influence on C, N, and P stoichiometry can also vertically occur in soil, showing different vertical variations along soil depth based on plant input, which depends on tree species. In addition, C:P and N:P can be predictors of carbon and nutrient cycles due to the strong relationship between soil enzyme activities and nutrient content [90]. Elemental stoichiometry can also be used as an indicator in monitoring urban forest vulnerability to nutrient imbalance along the atmospheric nitrogen deposition gradient. In atmospheric deposition, NH₄-N, PO₄-P, and NH₄:NO₃ were positively correlated to N and P contents in plant trees, litter, and soil [86]. In addition, stoichiometry can evaluate the efficiency of management practises in maintaining urban forest functions. The removal of understory vegetation is one of the standard management practises that can change the resource stoichiometry, leading to nutrient imbalance. The persistence of biodiverse-rich understory vegetation allows for avoiding changes in C:N, C:P, and N:P in both soil and microbial biomass [91].

The main studies reporting an ecological stoichiometry approach are shown in Table S1. However, to our knowledge, studies that have applied ecological stoichiometry for designing and monitoring NbS are lacking. The first attempt was carried out by Scartazza et al. [92], linking nutrient interaction and stoichiometry with physiological traits in a NbS created for air quality improvement. This approach contributes to selecting suitable species for NbS design, development, and management in urban areas.

3.2. Soil Enzyme Activities and Ecoenzyme Stoichiometry

The composition of plant species has an important influence on the properties of soil biochemistry [93]. Litter and root exudates provide the soil with organic matter, which includes starch, soluble sugars, organic acids, and amino acids. Additionally, plant cell walls provide cellulose, lignin, and biopolymers. The balance of macronutrients in soil surface horizons is affected by the type of litter and, thus, by how quickly organic matter breaks down [94]. Assessing soil biological indicators, such as enzyme activities, is essential to comprehending the interactions between living and non-living soil components, alongside biophysical and biochemical reactions. This comprehension aids in evaluating the impact on vegetation performance and overall soil health [95]. Because soil enzymes are linked to microorganisms and plant root activities, they are essential to the nutrition cycles. They are sensitive biomarkers of variations in the soil environment [96], microbial nutrient demand, and metabolic processes [97]. Enzymes such as β-Glucosidase and b-D-cellobiosidase are essential for the mineralisation of soil organic C. In contrast, phosphatase and N-acetyl-b-glucominidase are critical for the P and N cycles. Similarly, the S cycle requires the enzyme arylsulphatase.

Hence, soil enzyme activities have been applied as a valuable soil health indicator in several case studies in urban contexts (Table S2). Figure 4 reports the world distribution and number of case studies related to ecoenzyme stoichiometry.

Shan et al. [98] emphasised the importance of soil enzyme activities as vital biological indicators for assessing the fertility of urban forest soils. Enzymes such as hydrogen peroxidase, dehydrogenase, and alkaline phosphatase are significantly associated with soil physicochemical properties, including pH, organic carbon, and total nitrogen. These enzymes offer a more immediate and direct reflection of soil quality compared to chemical indicators, thereby serving as effective tools for evaluating the impact of soil management practises on soil quality and fertility. The different trends observed in the pattern of several enzyme activities were reported in the literature. Li et al. [99] highlighted a significant reduction in soil enzyme activities, such as dehydrogenase, catalase, and urease, in urban soils compared to rural soils along major roads in Beijing, China. This decline in enzyme activities indicates a degradation of soil quality, adversely affecting nutrient availability and plant growth. The lower enzyme activities suggest diminished microbial activity, essential for nutrient cycling and soil fertility. Cardelli et al. [100] aimed to evaluate the influence of the urban environment on soil quality by examining soil properties in 31 green areas within Pisa City, Italy. The research found that enzyme activities exhibited varying levels of variability; enzymes such as dehydrogenase and catalase showed little variability, whereas arylsulphatase and urease varied significantly. Castaldi et al. [101] investigated five sites with differing traffic densities in the urban area of Naples, Italy, while Teng et al. [102] estimated the potential ecological risks associated with vanadium in the soil of Panzhihua Urban Park, China. Both authors found that several microbial parameters, including dehydrogenase, sulphatase, glucosidase activities, and respiration, exhibited a significant decline following an exponential decay pattern as metal concentrations in the soil increased. Meanwhile, phosphatase activity and the average microbial response decreased in a sigmoidal manner. On the other hand, Hagmann et al. [103] highlighted that enzyme activities in urban soils subject to prolonged metal contamination can be unexpectedly high. In a study conducted in Jersey City (NJ, USA) analysing four historically contaminated soils, the authors found that soil microbial communities may adapt to elevated metal levels over time, resulting in increased extracellular enzyme function and nutrient turnover. Moreover, they emphasised the importance of considering urban soils’ historical context and specific environmental conditions when examining enzyme activities and their contributions to soil health and function. Jaworska and Lemanowicz et al. [104] assessed the impact of car traffic on the concentration of heavy metals in the park soil in Bydgoszcz (Poland). They explored the relationship between these metal contents and enzyme activity, reporting positive correlations between the presence of heavy metals, specifically lead (Pb) and cadmium (Cd), and the activity of various soil enzymes. A positive correlation between enzyme activities and organic carbon content was observed as expected. Bartkowiak et al. [105] emphasised the influence of proximity to a busy street in Bydgoszcz (Poland) on soil enzymatic activity, specifically noting a reduction in enzyme activities compared to a control area. Additionally, Gómez-Brandón et al. [106] discussed the role of enzymatic activities in urban soils as indicators of changes in soil quality, concentrating on a study conducted in Santiago de Compostela (Spain). They observed that urban soils exhibited higher activities of enzymes such as acid and alkaline phosphomonoesterases, which are sensitive to environmental conditions and changes in land use. The study underscored the potential of using enzymatic activities to monitor urban soil conditions, although it notes that land use only significantly influenced alkaline phosphomonoesterase activity. Gómez-Sagasti [107] applied enzyme activities as a valuable tool to evaluate soil health improvement in a phytoremediation study in urban soils in Vitoria-Gasteiz (Spain). Matinian et al. [108] highlighted that enzyme activity in the urban soils of Saint Petersburg (Russia) was affected by factors including organic carbon content, pH, and nutrient availability. Notably, soils from industrial zones exhibited the highest enzymatic activity due to optimal conditions for microbial growth. Interestingly, the expected negative impact on enzyme activity was not observed despite heavy metal pollution.

Ao et al. [109], in a research paper in an urban park in Beijing (China) on paired rhizosphere and bulk soils of the 12 woody species, revealed that root and soil nitrogen status influenced the rhizosphere effect on enzyme activity, thus highlighting the role of root functional traits in controlling soil nutrient cycling through the rhizosphere effect. Lemanowicz et al. [110] assessed how different tree species in urban areas affect soil quality and enzymatic activity, examining various enzymes such as dehydrogenases, catalase, and phosphatases while also investigating the risk of soil contamination by heavy metals like Pb, Ni, and Cd, considering selected physicochemical properties. The study was conducted in an urban park in Poland, and the authors found significant variations in enzyme activity patterns. Maienza et al. [111] applied the enzyme activities to highlight the significance of biological restoration and the fertility of urban soils following desealing in three urban locations (Carpi, San Lazzaro di Savena, and Forlì) in Italy. The study noted that soil enzyme activities, microbial biomass, and soil respiration displayed greater topsoil variation than desealed soils across various sites. Furthermore, ornamental shrubs were found to enhance soil organic carbon levels, likely supporting enzyme activity by providing substrates through leaf litter and rhizodeposition. Gorbov et al. [112] also examined how desealing influences the enzyme activity pattern in Rostov (Russia) urban soils. The research indicated that enzyme activities in urban soils are significantly lower than in their natural counterparts, with the most pronounced reductions seen in sealed and buried soils. The authors attributed this decline to altered soil formation conditions and a lack of organic matter migration. In a comparative study on urban soil health in two cities, Toruń (Poland) and Marrakech (Morocco), Berogui et al. [113] found that enzyme activities were affected by soil pH and organic matter content, with higher organic matter levels correlating with increased enzyme production. They also stated that enzyme activities, particularly dehydrogenase, alkaline phosphatase, and urease, serve as sensitive indicators of urban soil health and are responsive to anthropogenic disturbances. Igalavithana et al. [114], in a study conducted in Seoul (Republic of Korea), proposed an index based on five enzymes to evaluate soil quality, which was successfully validated using both unused soils and existing published data. The findings suggested that soils of high quality were associated with a high-valued index, providing a reliable method for assessing soil quality in urban agriculture based on the species present.

Several authors have recently recognised the importance of enzyme activities as indicators of urban soil health [115,116,117]. Enzyme activities are particularly highlighted for their ability to dynamically reflect changes in microbial community structures influenced by plant species and traits, making them valuable for evaluating interactions and overall health in urban ecosystems. Additionally, enzyme activities are recognised as reliable, sensitive, and informative indicators for assessing the decomposition of soil organic matter and a wide range of ESs in urban environments, supported by standardised methods for their determination. However, some disadvantages in evaluating urban soil health status through enzyme activities are foreseeable. The analyses are deemed costly and technically complex, requiring specialised equipment and expertise, which can impede widespread implementation. Urban soils are highly heterogeneous, comprising a mix of natural and anthropogenic materials, thus necessitating meticulous sampling to ensure accurate results. This heterogeneity can result in variability in enzyme activity measurements, complicating the interpretation of outcomes. Furthermore, pollutants, heavy metals, and other contaminants in urban soils can disrupt enzyme assays, potentially leading to misleading evaluations of soil health. These factors make it difficult to rely solely on enzyme activity for assessing urban soil health [115,116,117]. Some of these challenges could be overcome by applying the ecoenzymatic stoichiometry approach [118], which assumes that soil microorganisms’ metabolic and stoichiometric requirements are reflected in the production of enzymes targeting specific nutrient resources. This approach, which integrates enzymatic stoichiometry (enzymatic ratios) and the metabolic theories of ecology, has been widely used as a valuable tool for assessing microbial nutrient limitations in numerous investigations, some of which aimed at determining these limitations through large-scale assessments [118,119,120,121,122,123]. However, to our knowledge, few authors have applied it to evaluate soil health status in urban areas [87]. The authors highlighted the utility of ecoenzymatic stoichiometry as an indicator of soil functionality and its sensitivity to environmental changes along urbanisation gradients in Pisa and Livorno (Italy) urban areas. They revealed an imbalance in microbial carbon and nutrient acquisitions, with increased enzyme production for carbon acquisition in urban areas, indicating reduced microbial phosphorus limitation.

3.3. Carbon and Nitrogen-Stable Isotopes

Stable isotopes have been proposed as a promising tool in urban ecology [124] and as biomarkers of air pollution [125] and heavy metal contamination [126]. Commonly stable isotopes applied in urban studies include C, N, O, H, and S. However, recent advances in mass spectrometry (i.e., multiple-collector inductively coupled plasma mass spectrometry) now enable the measurement of heavier stable isotopes, such as Pb, Cu, Zn, and Se. However, the most frequent studies involve the analysis of C and N stable isotopes in plant and soil urban samples as suitable indicators for analysing the impact of urbanisation on C and N cycling. Therefore, for this review, the main publications on this topic were selected, classifying them based on the isotope form (δ¹³C, δ¹⁵N or both) and the ecosystem urban component (soil, vegetation, or both) investigated, as well as the location where the studies were carried out (Table S3, Figure 5).

3.3.1. Case Studies on Carbon Stable Isotopes in the Urban Soil–Plant System

Concerning C stable isotopes, rapid urbanisation resulted in increased emissions of CO₂ from fossil fuel combustion characterised by relatively low δ¹³C; hence, it has been hypothesised that plants growing in urban environments should be characterised by ¹³C-diluted values which, in turn, leads to a decrease in the δ¹³C values of organic matter entering soil C pools [127]. According to this hypothesis, a significant decrease in leaf δ¹³C values, associated with urban–periurban–rural gradients fossil CO₂ fuel emissions, was observed in grasses growing near a major highway in Paris [128], in Chinese pine growing in the Beijing metropolitan region [129], in leaf samples of Taraxacum officinalis growing in the Gent city areas [130], in urban red maple trees of three major cities arranged along a US latitudinal gradient [131], and in tree rings of Platanus hybrida in the urban area of Palermo (Italy) from 1880 to 1998 [132]. However, plant C isotopic signalling in urban areas can also be determined by factors other than CO₂ emissions, affecting C isotope discrimination during photosynthesis, such as microclimatic conditions and air pollutants. For example, plant δ¹³C in the Los Angeles Basin was significantly and negatively correlated with soil moisture and with concentrations of atmospheric pollutants [133], and the δ¹³C of the epiphytic CAM Tillandsia recurvata has been proposed as a useful indicator of atmospheric carbonaceous emissions in the Valley of Mexico basin [134]. Batkhuyag et al. [135] observed an increasing trend in tree-ring δ¹³C values over recent decades for all the investigated species in the Ulaanbaatar area in Mongolia due to increasing plant stress conditions, suggesting that climate change rather than air pollution affected the isotopic signature in plant urban species. The same increasing δ¹³C trend, associated with a consistent decrease in tree ring width, was observed also in Pinus pinea trees growing adjacent to main roads in the urban area of Caserta (South Italy) [136], and in five deciduous broadleaved tree species growing in neighbouring cities of Karlsruhe and Rheinstetten (Southwest Germany) [137], as a consequence of environmental stress.

The isotopic signature of urban soils depends on the source of C entering the soil C pool, which can be either ¹³C-enriched (e.g., inputs from C4 plants and calcium carbonate) or ¹³C-depleted (e.g., incorporation of fossil fuel-derived CO₂ into organic matter), and the underlying parent material [6,138]. The high contribution of carbonate-rich materials derived from anthropogenic sources (e.g., construction material, metallurgic activities) leads to a ¹³C enrichment in the total soil urban C [139]. Trammell et al. [6] found that total soil δ¹³C increased and organic soil δ¹³C decreased with increasing home age, indicating greater inorganic C sources in the yards around newer homes. Xu and Bai [140] showed that δ¹³C values in urban soil organic carbon in Shanghai (China) were related to the time since the conversion of soils to urban use. Conversely, Konstantinov et al. [141] found that δ¹³C values in the urban soils of Tyumen (China) were more related to the geological and landscape parameters of the study area rather than to urban disturbances and the development of urban ecosystems.

Overall, these studies highlight that C-stable isotopes in the soil–plant system can be used to study alterations in biogeochemical cycles and as reliable indicators of environmental constraints and ecosystem functionality in an urban context. The potential use of leaf δ¹³C represents another possible application of C-stable isotopes for improving urban ecosystem resilience as a quick and valuable tool for evaluating the water-use efficiency (WUE) of urban vegetation and for the selection of greening plants for future sponge cities characterised by high WUE [142,143].

3.3.2. Case Studies on Nitrogen Stable Isotopes in the Urban Soil–Plant System

Regarding N stable isotopes, several studies worldwide measured δ¹⁵N values in the urban soil–plant system as a powerful biomarker of atmospheric pollution and N cycling (Table S3). In particular, leaf δ¹⁵N has been proposed to indicate leaf uptake of atmospheric N pollution [144]. The distance to the road significantly influenced the δ¹⁵N values of tree leaves, suggesting a direct influence of vehicle NO₂ emissions on plant N isotopic signature [145,146,147,148,149,150,151]. Analogously, several other studies have observed an increase in δ¹⁵N in the leaves of trees in urban sites compared to those in periurban and rural areas [151,152,153,154,155,156]. These studies attributed the ¹⁵N enrichment in leaves to a greater contribution of NOx-N than NHx-N to urban N deposition along the urban–periurban–rural gradients. Conversely, a ¹⁵N depletion of plants and their epiphytes was observed in Brazil near industrial pollution sources dominated by isotopically depleted N forms such as NH₃ and NHx [157], while lichen C, N, and S contents and their stable isotope compositions (δ¹³C, δ¹⁵N, and δ³⁴S) depended on a complex mixture of airborne NOx and NHx compounds and anthropogenic sulphur sources in the city of Manchester [158]. Fang et al. [159] showed that leaf δ¹⁵N alone may not be a good indicator to assess the regional pattern of N saturation along an urban–rural transect in Southern China but that it can provide useful information on N status and N cycling in combination with measurements of soil N availability, leaf N concentration, and stream N export.

While the analysis of leaves has been applied to evaluate spatial changes in plant δ¹⁵N and anthropogenic N emissions, temporal changes have been investigated through isotope dendrochronology using tree rings. In this context, tree ring δ¹⁵N analysis, alone or in combination with other isotopes, has been applied as a biomarker of long-term changes in N deposition and pollution emissions in urban sites [135,146,160,161].

Several studies analysed δ¹⁵N changes in the soil–plant system of forests and lawns and urban–periurban–rural gradients by collecting leaf and soil samples. For example, similar δ¹⁵N trends in soil and leaf samples were observed in the soil–plant system of subtropical urban forests, with a ¹⁵N enrichment from the periphery to the centre of the Sao Paulo megacity due to NOx emissions in their respective surroundings [162]. A similar urban–periurban δ¹⁵N pattern was observed in soil and leaf samples in a Mediterranean urban city due to direct NO₂ deposition and indirect effects on the N cycling rate [87,154]. Konstantinov et al. [141] explained the spatial δ¹⁵N variation in urban soils of Tyumen (China) by changes in atmospheric N deposition. In contrast, the latitudinal trend of topsoil δ¹⁵N across a north–south transect in China has been explained by a combination of changes in climatic conditions (i.e., mean annual temperature and precipitation) and atmospheric depositions of both ammonium and nitrate [155]. Residential lawns in US cities had less variable and enriched plant and soil δ¹⁵N values relative to the associated native ecosystems across the major ecological biomes and climatic regions [6,163]. These results suggest that N sources to lawns, as well as greater rates of N cycling combined with subsequent N losses, may be important drivers of plant N dynamics in lawn ecosystems at the national scale, leading to δ¹⁵N convergence at the continental scale [6]. The greater soil δ¹⁵N in yards around older homes was attributed to a proportional increase in ¹⁵N enriched sources and accelerated soil N cycling rates associated with ¹⁴N removal over time, especially in poorly drained soils [164]. The soil–plant δ¹⁵N enrichment factor (i.e., the deviation of plant δ¹⁵N from soil δ¹⁵N) has been applied as a proxy of NO₂ deposition in Bromus spp. ecosystems of the Los Angeles Basin [133].

The effects of sociodemographic variables on the δ¹⁵N of plants and soil have been evaluated in association with urban expansion, housing age, and changes in farming practises [165,166,167,168,169]. Generally, most of the environmental changes associated with urbanisation (e.g., fragmentation, N deposition, changes in temperature and precipitation regime, housing age, fertilisation practises, etc.) lead to a ¹⁵N enrichment in the soil–plant system of the more urbanised forests and lawns relative to the pristine ecosystems [170].

3.3.3. Relationship Between C and N Stable Isotopes in Urban Contexts and Final Considerations

Studies combining C and N stable isotopes reported a positive relationship between δ¹³C and δ¹⁵N in organic soil in Mediterranean urban sites due to a combination of urbanisation degree, microclimatic conditions, and NO₂-induced air pollution, which simultaneously affected soil C and N cycling rates and the associated ¹²C and ¹⁴N losses over time [87,154]. A similar positive relationship between δ¹³C and δ¹⁵N in organic soil was observed in subtropical urban forests by Ramon et al. [56], suggesting that isotopic indicators can assess the suitability of urban soils as potential NbS, leading to benefits to the urban ecosystem. To summarise, all the above-mentioned studies highlight how C and N stable isotopes in the soil–plant system can effectively indicate urbanisation degree and related environmental and socio-economic factors that affect C and N biogeochemical cycles. Therefore, these indicators provide useful information on ecosystem functionality and NbS efficiency in an urban context. However, their interpretation is complicated by the many possible factors influencing their values in the extremely complex and heterogeneous urban environment and by the lack of reference values and standardised methodologies for the different plant and soil ecosystem components analysed.

4. Conclusions and Perspective

Research into implementing urban forestry to protect and restore ecosystem health and function has increased significantly over the past decade. In the light of this, many efforts are necessary to validate databases and procedures, guarantee scientific robustness, and provide policymakers and stakeholders with valuable tools for evaluating the effectiveness of established urban forestry. In this review, we highlighted the importance of the selection, validation, and measurement of harmonised soil–plant indicators for the effective planning and management of urban forestry, as well as for the evaluation of their performance in the long term. To this aim, we discussed the strengths and limitations of some soil–plant system indicators representative of nutrient cycles: ecological stoichiometry, enzyme activity and stoichiometry, and carbon and nitrogen stable isotopes. If evaluated together, this set of soil–plant parameters are effective indicators, providing a comprehensive picture of the impact on nutrient cycling and soil–plant health status in urban areas. However, given the heterogeneity and complexity of the urban environment, to promote their application to a wider scale, it is necessary to achieve the following: (1) increase our understanding of the environmental site-specific factors (e.g., climate, air and soil pollution, vegetation type, and soil heterogeneity) affecting these soil–plant indicators; (2) validate them across a range of urban contexts and spatio-temporal scales; (3) develop standardised procedures and establish reference values. This could enable selection of the most suitable site-specific methodological protocols to monitor the performance of urban forestry interventions, considering the complex heterarchical and bidirectional effects involving plant strategy and soil biota.