Formation, characteristics and eco-environmental implications of urban soils – A review

Abstract

Soils in urban areas have versatile functions, and their ecological services, especially the ability to buffer and purify pollutants, are very much needed. However, the formation and characteristics of urban soils are strongly affected by human activities, and so are their functions. In urban areas, soil horizons are often irregularly established, with anthropogenic layers and a high degree of heterogeneity owing to human disturbance and infill. Soil structure is often under degradation due to artifacts and technogenic substrata, mechanical compaction and human trampling. Physical degradation, such as compaction and destruction of structure, may reduce the ability of urban soils in infiltrating water and storing capacity, thus causing a higher runoff and pollutant load to the receiving water bodies. Soil compaction is also believed to worsen city heat island effects by reflecting more radiation, and seriously reduce the soil gas capacity and exchange. However, soil sealing fully loses its service function, further deteriorating urban ecology and environment. Enrichment of various waste materials associated with human activities, including nutrient elements, heavy metals and organic pollution, is the major problem of the urban soil environment. Materials containing rich liming in the urban soils lead to alkaline soil conditions, which can change element speciation and activities. The main features of urban soil contamination are characterized by strong accumulation of so-called “urban elements” such as copper (Cu), zinc (Zn), lead (Pb) and mercury (Hg), but less accumulation of other heavy metals. During the process of urban development, heavy metal contamination of urban soils not only happens nowadays, but also did in the past, especially when primitive mining and metal processing prevailed. Furthermore, urban soils are often polluted by organic pollutants, especially polycyclic aromatic hydrocarbons (PAHs). Platinum group elements (PGE), especially platinum (Pt), palladium (Pd) and rhodium (Rh) from automobile exhaust catalysts, are also accumulated in urban soils as affected by traffic. Physical and chemical degradation decrease soil enzymes and microbial activities, especially in industrial zones with high concentrations of heavy metals, low concentrations of organic matter and serious compaction. Unfortunately, the evolution of urban soil and attenuation of ecological function are believed to ultimately harm human welfare. Therefore, it is very important to establish a risk assessment system for better management of urban soil resources.

Key words:

• soil characteristics
• soil degradation
• soil environment
• soil formation
• urban soil

1. INTRODUCTION

More and more people are living in cities worldwide. In 2014, 54% of the global population lived in urban areas while in the most urbanized regions, such as Europe, the percentage was 73% (United Nations 2014). The primary impact of increasing urban population on urban environment is the mass generation and diffusion of urban wastes. Globally, more than 80% of urban and industrial wastes are from cities, which cover not more than 2% of the land surface (Zhang 2005). A big share of urban waste enters into urban and adjacent ecosystems, directly or indirectly. In many of the cities without adequate municipal waste treatment systems, urban soils instead are playing an important role in receiving and purifying the kinds of pollutants contained in solid, liquid and gaseous wastes. Under this circumstance, the ability of urban soils in providing eco-service is facing serious threats as the heavy environmental loading is using up the soil environmental capacity. In other words, soils in urban areas will probably shift from a pollution sink to a source, when their environmental capacity is exceeded, which may bring about a long-lasting environmental risk.

Since the 1990s, more and more concerns have been paid to urban soils globally (De Kimpe and Morel 2000), as urbanization has brought in remarkable physical, chemical and biological changes of urban soils. With native and agricultural lands shifting into urban environments, soil developmental trajectories are directly disturbed and even diverted. Human activities dominate soil formation and pedogenesis.

The potential hazard related to urban soil pollution is commonly understood as urban and suburban soils not only connect to human beings directly by touch and inhalation, but also affect human health indirectly through other exposure pathways such as the food chain. Urban water and air quality are also influenced by urban soil (CIDA 1998). Ecologically, urban soil is the living medium of urban plants and the inhabitance of soil microorganisms, as well as the sink of urban pollutants; therefore, urban soil has a pivotal role in balancing the material cycling of the urban ecosystem. The protection of urban soil itself is an effort to improve the management of the urban environment. Overall, the degradation of urban soil means the lowering of its ability and functions. The physical, chemical and biological processes as well as their impacts of urban soil degradation have become the focus of current urban soil studies.

2. FORMATION OF URBAN SOILS

2. 1. Impacts of human activities

Soil formation is an ecological process that results from the interactions of factors including climate, organisms, parent material, topography and time that weather parent material into soil mineral particles and add organic matter to soils (Jenny 1941). However, urban soils are located in urban and suburban areas and are strongly affected by various human activities; thus, their original nature is strongly changed (Effland and Pouyat 1997; Zhang et al. 2003b; Pickett and Cadenasso 2009). Commonly, humans influence soil properties and soil formation in an urban setting by construction, industrial pollution, burning of fossil fuels, waste inputs, fertilization and artificial filling (Craul 1992). As a result, human activities contribute novel anthropogenic parent materials to soils, change soil compositions and create strong heterogeneity.

Although natural pedological formation processes work simultaneously, human activities, especially physical disturbance, are the main driver for the formation of urban soils (Burghardt 1994). With urbanization, the original soils are often buried, truncated, removed and compacted. Furthermore, wastes from human activities are added to soil. Commonly, these foreign materials are all sorts of construction rubble, artifacts and building debris with cement, mortar, concrete and brick fragments (Stroganova and Agarkova 1993; Jim 1998a). In addition, there are also some chemical additives, mainly including heavy metals (Yang et al. 2011; Nannoni et al. 2014) and organic pollutants (Wilcke et al. 1999; Wang et al. 2013).

Besides physical disturbance, compaction is another important process. Because of various artificial forces associated with static loadings and dynamic impacts (Jim 1993), soils are often compacted in urban areas. In general, there are two types of compaction in urban areas: deliberate compaction during construction activities and unintentional compaction of the soil after construction is completed (Jim 1993, 1998a; Jim and Judith 2000). Deliberate compaction includes increasing strength for paving and housing foundations, the use of heavy equipment for reshaping and sloping banks along roads and hillsides and grading lots, and placing sod on hard soil or soil denuded of topsoil. Unintentional compaction includes allowing uncontrolled traffic (both vehicle and foot traffic) and allowing vehicles on lawn areas around homes or businesses, especially when the soil is wet (Yang et al. 2004b). When soil is compacted, the original aggregates are broken, united or rearranged (Jim 1993). Some interpedal and continuous pores are collapsed, and the elongated pores in the form of thin fissures are altered to parallel the soil surface associated with platy structure, so the air capacity and total porosity decrease (Jim 1986; Pagliai 1987). Following are changes of soil water, air and heat regimes (Yang et al. 2011; Huong and Pathirana 2013; Cheon et al. 2014).

2.2. Urban soil substrata

Urban soil substrata are more complex than parent materials of rural soils. The major natural substrata are refilled or redeposited, and usually are sandy with much gravel in urban areas (Jim 1998a). Besides original parent material or redeposited natural substrata, urban soils have a lot of technogenic substrata. Brick debris, rubble, ashes, slags, garbage, coal-mine deposits, bottles, batteries, plastics, metals, textiles, lino, charcoals and organic residues are common foreign materials in urban soils (Schleuss et al. 2000; Thiombiano and Gnakambari 2000). Soil layers often consist of these solid wastes (Lu et al. 2002). According to a recent study, in the city of Berlin, bricks can be found in all investigated soils (Nehls et al. 2013). Furthermore, the fine earth fractions contain 3–5% bricks, while the coarse fractions contain up to 50% (Nehls et al. 2013).

Some studies showed that urban soil contained obviously higher black carbon (BC) compared with agricultural soils. The BC might come from coal-mine deposits or ashes (Schleuss et al. 2000; He and Zhang 2006). Among various urban zones, roadside soil has a higher BC content and, meanwhile, a higher ratio of BC to total organic carbon (OC). Other zones, such as parks, residential quarters and school yards vary in BC content, which indicates the diversified sources of BC (He and Zhang 2006).

The special technogenic substrata could appear in the whole soil profile to more than 6 m (Schleuss et al. 2000; Zhang et al. 2007). Burghardt (2000) classified the substrate according to the method of deposition of material, the kind of man-influenced and man-made material, the modification of the material by treatment and the content of skeletons irrespective of fine earth. Those technogenic substrata not only reflected the history of human activities, but also helped with the soil classification.

2.3. Record of past human activities in urban areas

Soils in urban areas are often artificially formed and preserved as cultural layers. The formation of so-called cultural layers has special soil features, usually mixed with artifacts. The vertical change of soil properties denotes the historical change of human activities, and, therefore, urban soil can be considered as a record of human activities and the history of urban development. Bryant and Davidson (1996) discussed the cultivation history of north Scotland in the 18th and 19th centuries with old cultivated soils. Zhang et al. (2005b) studied a 6-m-deep profile consisting of cultural layers of different dynasties, which recorded the entire deposition history during the past 20 centuries in Nanjing, China. Zhang et al. (2007) found, using a stable lead (Pb) isotope technique, that within ancient urban soil layers, Pb and other heavy metals increased substantially due to metallurgical activities. Some other investigations in old cities such as Moscow, Russia and Constanz, Germany (Stroganova et al. 1998; Roeber et al. 2000), included also the buried part of urban soils and attributed the abnormal amount of trace elements to past human activities and land uses. Roeber et al. (2000) found that primitive industrial processing of metals in the Middle Ages caused the accumulation of copper (Cu) and Pb in subsurface soils. Therefore, the formation and evolution of urban soil usually record some special materials due to human activities with different historical characteristics.

3. MAJOR CHARACTERISTICS OF URBAN SOILS

3.1. Physical characteristics

3.1.1. Appearance of coarse fragments

The presence of coarse fragments in urban soils is ubiquitous, as a result of artificial transportation of natural rocks and, more often, the incorporation of artificially made materials such as concrete and building bricks (Jim 1998a). The stone and sand contents are commonly very high (Short et al. 1986; Greinert 2000; Jim 2001). The stone content ranges between 13.4 and 81.8%, and the sand content between 71.9 and 92.1% in Hong Kong, China (Jim 2001). Studies showed that texture is only partly inherited from the siltstone parent material, or found no discernible effect of parent material influence (Jim 2003). The disturbance has increased the coarse fractions (stone and sand), and decreased the fine fractions (silt and clay). The difference between less disturbance and more disturbance is stronger when stone contents are more than 20%, and especially when they are more than 40% (Jim 2003). To some uncompacted sites, the change of soil texture favors the preferential flow of solute and the transportation of pollutants, which means a more direct influence to groundwater as soil does not fully function as a filter in this case. To the compacted sites, usually surface compaction is more serious than subsurface (Yang et al. 2004b). Therefore, although urban soil contains coarse fragments, the infiltration rate is often low due to heavy compaction (Yang and Zhang 2011).

3.1.2. Soil compaction and sealing

Another common soil physical characteristic in urban areas is compaction. Compaction is the most serious form of physical degradation, which is basically the reduction in volume of a given mass of soil with an increase in soil bulk density, closer packing of solid particles and decreased porosity (Glinski and Lipiec 1990; Jim 1993; Yang et al. 2004b). Data from Nanjing, China, showed that soil bulk density varied from 1.14 to 1.70 g cm⁻³, mostly higher than the ideal value for plant growth (Yang et al. 2005). Jim (1998b) also reported that in Hong Kong, China, soil bulk density in urban street shoulders varied from 1.14 to 2.63 g cm⁻³ with an average of 1.67 g cm⁻³, and the average total porosity of soil transported into city for tree planting was 36.6% (24.8–59.2%; Jim and Judith 2000) showing serious compaction.

Other soil physical properties are strongly influenced by compaction, which includes the destruction of soil structure, loss of total and air porosity, lowering of infiltration rate and saturated conductivity, and decrease of effective water content and water-adjusting ability (Yang et al. 2006). Compaction also means the increase of soil strength, hence a higher resistance to the root of urban plants (Jim 1998a). Moreover, soil compaction tends to minimize the contact of soil particles with water and leads to a slow diffusion of oxygen (O₂) and carbon dioxide (CO₂). As a result of the above-mentioned processes, soil compaction has many detrimental impacts on the urban ecosystem, which includes: (1) reduced water infiltration and recharge to groundwater system; (2) increased surface runoff and promoting the formation of urban flooding; (3) increased pollutant loading in surface water bodies; (4) increased urban heat island effect; (5) affected soil temperature, microorganism activity and nutrient transformation; (6) affected plant growth by weakening root activity and its biomass thus shortening plant longevity. As a whole, urban soil compaction can fundamentally affect the urban environment and likely lower the living quality of the urban population.

Sealing by urban gray infrastructure, which includes all forms of pavements and buildings (Breuste 2011) – i.e., impervious surfaces, such as asphalt and concrete – is very common in urban areas. Data showed that 2.3% of the European territory was sealed (Prokop et al. 2011), and 5% of the German territory was covered by impervious surfaces (Prokop et al. 2011). Estimates showed that 46–50% of transport and settlement areas were sealed (Breitenfeld 2009). Sealing has especially negative impacts on the potential provision of ecosystem services. For example, it hampers the exchange of material and energy between the soil and other environmental compartments, increases water surface runoff (Haase and Nuissl 2007) and interferes with microclimate regulation by increasing temperatures (Henry and Dicks 1987).

3.1.3. Degradation of soil structure

Soil structure is poor, with less aggregation in the urban region. The shortage of fine particles limits the formation of a strong soil structure due to the lack of aggregating agents and bridging materials between coarse grains. Soil compaction resulting from urbanization can alter soil aggregate arrangement and pore space. The low organic material and high stone and sand contents are reasons for low aggregates because organic material is a significant component of the binding agents that form aggregates (Six et al. 2004). Urban land development processes may involve vegetation clearance, topsoil removal, stockpiling, compaction, building or soil replacement, all of which can drastically affect soil aggregation (Wick et al. 2009). For example, topsoil removal and replacement resulted in about 29% average reduction in the proportion of macroaggregates in the surface soils (Chen et al. 2014). Jim (1998a) found that the proportion of water-stable aggregates in highly disturbed roadside soils in Hong Kong was very low.

Urban soils fail to develop the granular or crumb structures normally developed in natural and agricultural topsoil. The single grain sites are tied to the sandy texture and to some artifacts such as ash, furnace slag (Jim 1998a; Yuan 2006). The common structure is an angular blocky and platy structure (Jim 1998a). Compaction could trigger a negative feedback, with compacted soil becoming more resistant to further structural degradation. During the compaction process, the vertical force has disintegrated the original structure and remolded materials to match the new stress regime, with particles and pores rearranged to lie perpendicular to the impact direction. In the urban park of Hong Kong, intensive trampling had re-organized particles down more than 10 cm. Some unnatural massive structures, often found in the urban subsoil (Yuan 2006), are probably an inherited feature formed by heavy-machinery compaction and earth stockpiling during construction of the park. The feeble structure-forming process has failed to rejuvenate the damaged layers.

3.2. Chemical characteristics

3.2.1. pH

The urban soil pH is slightly to very strongly alkaline, according to the majority of studies. As reported, soil pH varied from 7.4 to 8.6 with most of values in the narrow range 8.0–8.2 in Siena municipality, central Italy (Nannoni et al. 2014). The roadside soils in Hong Kong, China, had a mean pH of 8.7, in which over half of the samples were rated strongly alkaline (pH 8.5–9.0) and very strongly alkaline (pH 9.0–9.5; Jim 1998c). In Szeged, Hungary, the urban soil pH was 7.6–9.1, while local hill soil developed on granitic material had a pH of 4.0–5.0 (Puskás and Farsang 2009). A study in Nanjing, China, showed that the pH gradient from rural to urban gradually increased from about 6.0 to more than 7.5 (Zhao et al. 2007).

The common high soil pH in urban areas is attributed to filling materials contaminated with building wastes, such as concrete and cement, with release calcareous solutions. The soil samples from the Mall in Washington, DC, containing filling materials of building rubble, had alkaline reactions (Short et al. 1986). In four urban parks in Murcia City (southeast Spain), all soils were alkaline with a pH range from 7.2 to 8.2, and calcium carbonate (CaCO₃) equivalent contents varied from 31% in rhizosphere soil to 59% in playground soils (Acosta et al. 2014). Puskás and Farsang (2009) showed a close correlation between pH values and carbonate content, and a congruent fluctuation of them in the studied profiles. Nannoni et al. (2014) had similar results in the Sienaurban area, Italy.

The alkaline soils in urban areas not only induce nutrient imbalance but also change the metal speciation and activities. For example, a study at three inactive railway yards on the Island of Montréal showed the activities of free metals [cadmium (Cd), Cu, nickel (Ni), Pb and zinc (Zn)] were very low, and those heavy metals were largely immobilized at the soil particle surface because of high pH or organic matter content (Ge et al. 2000). Furthermore, over 40% of Ni was present as inorganic complexes if the solution pH exceeded 8.1 (Ge et al. 2000). Janssen et al. (1997) even found a significant effect of pH on Cd, Pb and Zn contents in earthworms living in the contaminated soils.

3.2.2. Phosphorus accumulation

Phosphorus (P) is an element needed by all life forms. The basic pattern of P cycling in urban ecosystems consists of the following pathways: the input of P through industrial and human uptake such as daily diet, which comes from food products and ultimately from agricultural soil; the export of P after human or industrial metabolism into sewage and waste products; the re-entering of P into urban and suburban soil as organic manure or sewage sludge; and the output of P into the urban water system from sewage water and soil release. The elements accumulated in urban ecosystems generally discharge in two major ways, i.e., the export out of cities by artificial and natural transportation, such as waste disposal and sewage water discharge, and the accumulation in urban soil or at least a very slow turnover in soil. The above cycling process is actually a general pattern for most of the life elements in the urban ecosystem. In the past 50 years, about 1 billion tons of P has been mined globally, of which 80% was used in agricultural land (Zhang 2005). However, globally, the increase of P in agricultural soil is not remarkable and it is believed that most of the applied P has been harvested. It is estimated that 250 million tons of P did not recycle back to agricultural soil and a large part of it remains in the urban environment, as nearly 40% of people live in cities and human consumption and excretion accounts for it accordingly (Zhang 2005). Considering the small percentage (less than 2%) of global land surface used by cities, the storage of P in the urban environment should be very high.

A study in Nanjing, China, showed that the total P content in urban soil could reach 15 g kg⁻¹, about 10 times that of normal agricultural soils. The amount of 0.5 mol L⁻¹ sodium bicarbonate (NaHCO₃)-extractable P was as high as 400 mg kg⁻¹, with an even higher fold. Except for those newly introduced soil materials in the urban environment, the major part of urban soil had a significantly higher P concentration. The build-up process actually began as soon as a city was established (Zhang et al. 2001), because the recycling of P with human waste for urban agriculture was a common practice in the past. There is normally a gradient distribution of soil P concentration in the urban–suburban transect, suggesting the spatial diffusion of P accumulation (Zhang 2004). High soil P concentration poses a potential hazard to the environment as the release of P to water can be rapidly enhanced (Zhang et al. 2005a).

As a matter of fact, nutrient elements other than P also are accumulated in urban soil. For example, nitrate usually accumulates in urban and suburban soils. Although the residence time of nitrate in soil is relatively short, the continuous input through the application or addition of organic manure or waste keeps it at a high level in urban and suburban soils. In addition, salt accumulation is also often seen in the urban environment, due to artificial addition by various activities. In a word, urban and suburban soils are usually “eutrophicated” (Zhang et al. 2006).

3.2.3. Accumulation and contamination of heavy metals

Heavy metals in urban soil come from parent material and exogenous input. Numerous studies have shown that heavy metals in urban soil are more concentrated than in surrounding agricultural and forestry soils, due to the strengthened addition of heavy metals during urbanization (Tiller 1992; Chen et al. 1997; Stroganova et al. 1998; Manta et al. 2002; Lu et al. 2003). A study in Siena municipality (central Italy) showed that soil Cd, Cu, Pb, antimony (Sb) and Zn contents increased in the following order: non-urban < green-urban < peri-urban < urban soils (Nannoni et al. 2014). Ding and Hu (2014) evaluated the peri-urban soil and showed that the Cd and arsenic (As) in the soil could result in high ecological risks.

The sources of exogenous heavy metals include household, garbage disposal, transportation, mining and processing, manufacture, power plant and other fossil fuel burning (Tiller 1992; Stroganova et al. 1998; Wong et al. 2006). Among those sources, combustion of leaded gasoline contributed the dominant part of Pb accumulation. Studies also showed that the abrasion of Zn-added tires produced Zn-containing particles which finally entered into urban soil (Liu et al. 1996; Yang et al. 2011). Traffic is believed to be one of the main sources of Cu, Zn and Pb in urban soil (Sáňka et al. 1995; Imperato et al. 2003; Yang et al. 2011; Nannoni et al. 2014). Cicchella et al. (2008) applied Pb isotope investigations and also demonstrated that road traffic was one of the main sources of metal pollution. Mercury (Hg) and As primarily originated from coal combustion, while Cd was mainly associated with industrial sources (Li et al. 2001; Yang et al. 2011).

The sources and concentrations of heavy metals in urban soil of different cities may vary according to industry types. Lu et al. (2003) found that in urban Nanjing, China, Ni, cobolt (Co) and vanadium (V) were mainly inherited from parent soil material, but Cu, Zn, Pb and Cd increased greatly by artificial input. Studies in Bangkok, Thailand (Wilcke et al. 1998), and Uberlândia, Brazil (Wilcke et al. 1999), also showed that soil iron (Fe), Ni, Co and V came from parent materials, while Cu, Zn, Pb and chromium (Cr) increased by human activities (Wilcke et al. 1998, 1999). In Palermo, Italy, it was found that Pb, Zn, Cu and Hg contents in urban green belt and parks were 4.6, 1.1, 1.9 and 9.7 times higher than background values, and statistics indicated that Pb, Zn, Cu, Sb and Hg were affected by exogenous contamination while manganese (Mn), Ni, Co, Cr and V were mainly from parent soil (Manta et al. 2002). Normally, the non-residual fraction of heavy metals in urban soil is higher, indicating a higher bio-availability (Lu et al. 2003). In general, most of the studies show that Cu, Zn, Pb and Hg are the main contaminating elements and they are typical “urban heavy metals,” while the accumulation of other heavy metal elements is not significant.

Soil contamination by heavy metals in the surface represents only one aspect of soil pollution in urban area. It is understood that surface soil contamination is caused by modern traffic and industrial activities. However, it is seldom recognized that historical development can cause soil pollution too, which often appears in the deep-buried horizons and becomes an important record of historical metal processing during city development (Alexandrovskaya and Alexandrovskiy 2000; Zhang et al. 2005b). It can be concluded that as soon as ancient industries appeared in urban areas, remarkable soil contamination started, but only recently was this recognized by urban soil studies (Yang et al. 2004a). As a matter of fact, urban soil has a similarity with solid sludge, which contains large amounts of plant nutrients such as P, and also considerable heavy metals, especially Cu, Zn and Pb. During the urbanization process, a massive amount of “old” urban soil is often replaced and disposed of without necessary soil quality evaluation, which may lead to a secondary off-site release of contaminants if not properly managed.

3.2.4. Accumulation and contamination of platinum group elements

Platinum group elements (PGE), which include platinum (Pt), palladium (Pd), osmium (Os), iridium (Ir), ruthenium (Ru) and rhodium (Rh), are valuable rare elements, typically present at concentrations of 0.4–5 μg kg⁻¹ in the earth’s crust (Wedepohl 1995). Their chemical characteristics make them highly useful as catalysts in a variety of chemical and pharmaceutical processes, especially as automobile exhaust catalysts (Moldovan et al. 2002). Some studies have shown that PGE are bioavailable in the environment and pose health risks at chronic levels (Ravindra et al. 2004). The PGE contaminants in urban soil mainly are Pt, Pd and Rh from automobile exhaust catalysts (Mihaljevič et al. 2013).

PGE inputs are greatest in urban areas, with the major sinks being road sediments and roadside soils (Sutherland 2003; Jackson et al. 2007; Sutherland et al. 2007; Spaziania et al. 2008; Mihaljevič et al. 2013). The maximum PGE concentrations were for Pt (160 μg kg⁻¹), with lower contents of Pd (49 μg kg⁻¹), followed by Rh (3.9 μg kg⁻¹), in Prague (Mihaljevič et al. 2013). In Hong Kong urban soils, the values were 160 μg kg⁻¹ for Pt, 107 μg kg⁻¹ for Pd and 34.5 μg kg⁻¹ for Rh (Pan et al. 2009). Higher concentrations of Pt (506 μg kg⁻¹), Pd (105 μg kg⁻¹) and Rh (64 μg kg⁻¹) were found in residential streets in Hawaii, and maximum enrichment ratios exceeded 400 (Sutherland et al. 2007).

PGE concentrations in the soil are affected by traffic. Firstly, there is a tendency for urban soil PGE concentrations to increase over time. In Braunschweig, Germany, a comparison of 2005 with 1999 revealed a distinct increase of PGE concentrations in soils close along heavy traffic roads, by a factor of 2.1–15 (Wichmann et al. 2007). Pt concentrations in urban soil collected in 2001 were up to six times more than those measured in 1992 samples in Rome (Cinti et al. 2002). A similar increasing tendency was found in the metropolitan area of Mexico City from 1991 to 2003 in the fine particulate matter (PM2.5 and PM10) (Morton-Bermea et al. 2014). Secondly, the soil PGE concentrations increased with a decrease in the traffic speed. For example, maximum concentrations in soil were 50.4 μg kg⁻¹ for Pt, 43.3 μg kg⁻¹ for Pd, and 10.7 μg kg⁻¹ for Rh beside a road with a constant speed of about 80 km h⁻¹, and were 88.9 μg kg⁻¹ (Pt), 77.8 μg kg⁻¹ (Pd), and 17.6 μg kg⁻¹ (Rh) beside a road with a constant speed of 50 km h⁻¹, while they were 261 μg kg⁻¹ for Pt, 124 μg kg⁻¹ for Pd and 38.9 μg kg⁻¹ for Rh with stop-and-go traffic in Braunschweig, Germany (Wichmann et al. 2007). Furthermore, soil PGE concentrations were the highest close to the road and declined with growing distance (Wichmann et al. 2007; Spaziania et al. 2008).

High PGE concentrations were detected especially in topsoil layers in an urban area (Wichmann et al. 2007; Mihaljevič et al. 2013). Sutherland (2003) found significant enrichment with depth in several roadside sites in Honolulu, Oahu, Hawaii, which may indicate potential Pt mobility in this environment (Sutherland 2003). Mihaljevič et al. (2013) estimated the velocity of migration of PGE through the profile corresponded to 1.1–2.2 cm year⁻¹. However, Pt concentration in upper soil horizons in roadside areas is significantly lower than in road dust (Sutherland 2003; Tsogas et al. 2009). This implies that PGE in urban soil mainly comes from road dust. Wichmann et al. (2007) monitored the airborne dust at roadside and revealed astonishing high values for Pt 1730 μg kg⁻¹, for Pd 410 μg kg⁻¹, and for Rh 110 μg kg⁻¹. However, PGE concentrations in rural airborne particulate matter are orders of magnitude lower than in urban aerosols (Sen et al. 2013).

3.2.5. Organic pollutants in urban soil

Both persistent organic pollutants (POPs) and persistent toxic substances (PTS) are volatile or semi-volatile organic molecules (Tang 2003), which can migrate into air or water bodies through volatilization, leaching or diffusion due to a concentration gradient, and are a serious hazard to air and water quality, and ecosystem and human health (Xue et al. 2002). POPs and PTS include polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and polychlorinated naphthalenes (PCNs). Human activity is the main source of environmental PAHs. Tang et al. (2005) suggested that PAHs of urban soil mainly come from pyrogenic sources according to the molecular component, while other studies indicated that traffic emissions might be the major origin of PAHs in soils (Wilcke et al. 1999; Wang et al. 2013; Yu et al. 2014). In fact, there are many sources of POPs in urban soils, such as petrogenic sources, coal combustion, biomass burning, creosote, coke tar-related sources and vehicular emissions (Wang et al. 2013). Soil is an important sink of PAHs. Over the past 100–150 years, the concentration of PAHs in soils has increased continuously, especially in urban areas (Ding and Luo 2001). Because of PAHs entering into soil by mainly atmospheric deposition and waste disposal, PAH concentration in soil depends on firstly its concentration in atmosphere or waste, and secondly soil properties affecting its transformation and movement in soil. As an important industrial material, PCBs are also sunk in soil; for instance, PCBs contained in soil account for 93.1% of their total in terrestrial ecosystems in the United Kingdom (Harrad et al. 1994).

Krauss and Wilcke (2003) found that PAHs, PCBs and PCNs in urban soils of residential green areas and industrial sites were several times higher than in rural agricultural soil, and showed a gradient decrease from the center to the outskirts. A study in Delhi, India, showed that PAHs in urban soil were seriously over (4–15 times) the national standard (Khillare et al. 2014). Toxicity potentials of those soils were 8–88 times higher than reference levels. A recent study in Hangzhou, China, showed that total PAHs had a surprisingly high concentration in urban soils, with a range of 181 to 1981 µg kg⁻¹ and a mean of 611 µg kg⁻¹ (Yu et al. 2014). Among different functional zones, a higher level of PAHs was found in the roadsides, followed by commercial districts, residential areas, parks and greenbelts (Yu et al. 2014). The generation and migration of organic pollutants in environment determine to a great extent the direction and range of their spatial concentration autocorrelation (Khillare et al. 2014). Total concentrations of PAHs in soils were strongly correlated with soil organic carbon (SOC), implying that SOC was the key factor determining the retention of PAHs in soils (Tang et al. 2005; Wang et al. 2013).

3.3. Biological characteristics

3.3.1. Biodiversity

The effects of urban sprawl on biodiversity are poorly known. Savard et al. (2000) put forward biodiversity concepts in urban ecosystems. Mcdonald et al. (2008) showed that urban growth would have impacts on ecoregions, rare species, and protected areas. They also showed that 29 of the world’s 825 ecoregions have over one-third of their area urbanized and 8% of terrestrial vertebrate species on the International Union for Conservation of Nature Red List are imperiled largely because of urban development. Therefore, urbanization is one of the leading causes of species extinction (McKinney 2006), not only in the developed countries such as United States (Czech et al. 2000), but also in less developed countries, for example Concepción, Chile (Pauchard et al. 2006). Native ecosystems are replaced by pavements and buildings, and what is left of the natural soil is covered with green areas dominated by non-native ornamental species. With the decrease of localized native species, non-native species will increase, which promotes biotic homogenization (McKinney 2006). Therefore, the massive disturbances created by city growth destroy the habitat of native species, while they create habitat for a relatively few species that are adapting to urban and suburban conditions (Sax and Gaines 2003; McKinney 2006; Mcdonald et al. 2008). The consequence for conservation is that non-native species may often enrich local biodiversity (Sax and Gaines 2003), but global diversity is decreased by the subsequent extinction of unique local species that are lost to the global species pool (McKinney 2006).

3.3.2. Activities of biota

Urban soil contamination would endanger living organisms. The study in Siena municipality (central Italy) showed that soil contamination influenced the uptake of Cd, Cu, Pb, Sb and Zn by earthworms, and there were significant positive correlations between Cd, Pb and Sb earthworm concentrations and their soil contents (Nannoni et al. 2014).

Urban soil pollution is also evidenced to cause changes of microorganism features. A study in Aberdeen, United Kingdom, found that the microbial base respiration ratio in urban soil was enhanced but the biomass significantly lowered, as compared with neighboring agricultural soil. The result showed that in urban soil, microbial consumption of OC was significantly increased (Yang et al. 2001). Further study through principal component analysis indicated that extractable soil Pb was the main factor determining the difference in microbial features between urban and agricultural soils, with available and organic-bound Zn, Cu and Ni as minor affecting factors. Compared with natural soils of the same region, urban soil in Nanjing, China, had a wider range of carbon (C)/N and microbe C/OC (C_mic/C_org) ratios, suggesting a remarkable interference of human activities with soil microbes in the urban environment (Wang et al. 2005). In Moscow, Russia, the portion of fungi decreased abruptly (by 2.5 times) in the soil of the industrial zone compared to those of the recreational zone and the natural analogue (forest; Ivashchenko et al. 2014). Furthermore, there is a tendency toward the deterioration of the functioning of the microbial community under anthropogenic transformation of the soil. Possible reasons for this are the soil compaction and contamination, and trampling may destroy the environment of the microbial community because major microbial activity (69–85%) is concentrated in the upper 10-cm-thick mineral layer (Ivashchenko et al. 2014). Urban sealing also results in a decrease in soil microbial activity and functional diversity, but the influences on soil microbial diversity vary among land uses, with road pavement having the most negative effect (Wei et al. 2013). In addition, microbial biomass increases along the depth gradient (Zhao et al. 2012). Soil chemical characteristics also affect microbial activity. The soil nitrogen (N) and P levels in the urban area have strong effects on enzyme activities and microbial processing of soil organic matter (Cusack 2013).

The intensity of the anthropogenic impact determines soil enzymes activities in urban soils. Bielińska et al. (2013) found that in cities of Poland and Upper Silesia, the activity of enzymes (dehydrogenases, acid phosphatase, alkaline phosphatase, protease and urease) was several times lower in the soils of gardens located in the city center than in the soils of gardens situated in the suburbs. Generally, the high concentration of heavy metals, low concentration of organic matter and high pH (alkaline) would decrease enzyme activities (Xian et al. 2014). The most serious disturbance and soil sealing in the urban areas also significantly affected enzyme activities (Zhao et al. 2012). In some circumstances, the activities of enzymes will increase. For example, in the 0–20 cm soil layer of green land irrigated with reclaimed water with enhanced OC, N and P concentrations, they were enhanced by an average of 36.7%, while they decreased with increasing soil depth (Pan et al. 2012). Therefore, soil biological activities could be improved with long-term reclaimed water irrigation.

3.4. Soil horizon development

Urban soils represent typical anthropogenic soils and exhibit a high degree of heterogeneity, with great vertical and spatial variability (Effland and Pouyat 1997; Baumgartl 1998; De Kimpe and Morel 2000). Very dense material intermixes with coarse substrates; layers of clayey soil are next to layers of loose ashes (Baumgartl 1998). The modification of original soil parameters and characteristics is often observed, including an elevated amount of artifacts, atypical soil structure, compaction, restricted aeration and water drainage, the presence of surface crusts on bare soil, modified soil reaction, interrupted nutrient cycling, modified soil organism activity, the presence of contaminants, and highly modified soil temperature and moisture regimes (Craul 1985; Puskás and Farsang 2009). Therefore, the soil physical and chemical properties of a considered volume are very heterogeneous.

In the urban soil profile, some soil horizons are missed or doubled, even overthrown. Some horizons fully consist of brick, mortar debris, residues of charcoal, ashes, rubble and so on. A typical feature of urban ecosystems is anthropogenic deposition of different materials several meters thick (Billwitz and Breuste 1980). The remainder of the old town is covered by Cumuli- and Rudi-Urbic Anthrosols mainly developed from mixtures of garbage and rubbles with sandy materials (Schleuß et al. 1998). In most natural soils, there is a gradual downward decrease of the humus content (Lorenz and Kandeler 2005), while in urban soils the humus-enriched layer was either stripped or buried. So there isn’t a normal regular content of humus in the soil profiles. Moreover, the majority of the profiles fall into the low humus content (1–2%; Puskás and Farsang, 2009) because of artificial landfill with low humus content in the parent material. The majority are mixed profiles embedding both infilled material and buried soil horizons (Puskás and Farsang 2009). At present, the main soil-forming processes are organic matter accumulation in and slight loss of alkalis downward from the A-horizon.

The anthropogenic horizons are also characterized by very high carbonate values with a gradual downward increase towards the parent material. The reason for this is the leaching of carbonate phases from the upper soil horizons, and the accumulation of these in the underlying layers or the parent material itself (Puskás and Farsang 2009). Therefore, the calcic horizon often exists in the urban soil. Another characteristic soil-forming process is clay translocation downwards into the B horizon (Schleuß et al. 1998) in more acid environments.

Some layer boundaries are abrupt, whereas others are gradual and diffuse. Sharp composition and property differences between layers usually suggest an artificial rather than a natural origin (Jim 1998a). A haphazard filling process with batch differences in materials could have induced layering. Filling and drastic disturbance also create soils with evident discontinuities in the profile (Weber et al. 1984; Short et al. 1986). The poor initial state of the parent material, compounded by continual trampling and associated erosion, has overwhelming impacts on soil development.

4. ECO-ENVIRONMENTAL IMPLICATIONS OF URBAN SOIL DEGRADATION

4.1. Water movement

4.1.1. Infiltration rate

Compaction reduces soil porosity (Glinski and Lipiec 1990) and increases soil bulk density; thus, soil water movement is affected by compaction. Yang and Zhang (2011) investigated infiltration rates with wide variance from less than 1 to 679 mm h⁻¹ in the residential areas, park areas, road greenbelt, and campus in Nanjing, China. Although the majority of infiltration rates were 5–63 mm h⁻¹, those less than 5 mm h⁻¹ accounted for 17% of the total determined sites. These low infiltration rates appeared in the severe and extreme soil compaction. Gregory et al. (2006) measured the infiltration rates on sandy soils in North Central Florida on urban construction sites and across various levels of compaction and found the infiltration rates on compacted soils ranged from 8 to 188 mm h⁻¹ for compacted sites, while non-compacted infiltration rates ranged from 255 to 652 mm h⁻¹ for forest and pasture. So construction activity or compaction treatments reduced infiltration rates by 70–99%. Statistical analysis found that infiltration rates were related to the bulk density, total porosity, air-filled porosity, capillary porosity and organic matter and sand, silt and clay contents, and concluded that soil compaction and soil texture were the primary factors affecting the infiltration rates (Yang and Zhang 2011). Winzig (2000) demonstrated that the bulk density and infiltration rate of a soil were highly correlated.

4.1.2. Flooding risk

Soil has a large storage capacity for water retention, which can supply water to plants for their growth and minimize flooding risk. However, soil compaction and sealing reduce or eliminate the abilities of soil holding and adjusting water balance. In urban areas, with the increase of soil compaction, the total soil water storage capacity, available water storage capacity and short-term water storage capacity reduced gradually, but the ineffective water storage capacity increased gradually (Yang and Zhang 2008). Compaction can weaken the soil environmental buffering function which is very important to the urban eco-environment.

As the ability of urban soil in discharging and holding water is reduced, not only is plant growth affected, but also rapid urban flooding often occurs. A reduction in the infiltration rate increases the runoff coefficient and then the increase of surface runoff by stormwater. As a result, the frequency of floods in urban areas with compacted soils is higher than in areas with non-compacted soils. The frequency of surface runoff for severely compacted soils was equal to 8%, while for extremely compacted soil, it was approximately 40% in the subtropical Nanjing City, China (Yang and Zhang 2011). Wang et al. (2000) studied the effect of compaction on the runoff coefficient by simulating rainfall, and found that the volume of runoff from compacted soil was 2.2 times greater than that of non-compacted soils. Existing impervious surfaces in urban areas will further add to the probabilities of flood occurrence. Haase and Nuissl (2007) found that surface runoff had more than doubled in the city area between 1940 and 2003 in Leipzig, Germany, due to the increase of impervious surfaces. This implies that in urban regions, the potential for runoff and the need for large stormwater conveyance networks are not only due to the increase in associated impervious area but also to the compacted pervious area.

4.1.3. Water quality

Urban soil can act as a water purifier by filtering and absorbing pollutants of surface runoff, but at the same time it can be a contamination source to water bodies as the considerable amount of sunk pollutants are subject to loss by various pathways. It is therefore necessary to know the environmental capacity of urban soil for the assessment of its environmental risks. A study about the soil adsorption–desorption of P in Nanjing showed that when Olsen-P was higher than 25 mg kg⁻¹, a rapid increase of P release from soil to water was observed (Zhang et al. 2005a), while the average Olsen-P content was about 64 mg kg⁻¹ which is believed to be a threat to surface- and groundwater quality (Zhang et al. 2001, 2003a). Evidently, N, P and some heavy metals of groundwater samples were enriched in the studied sites in Nanjing.

Together with runoff, concentrated pollutants such as ammonium, heavy metals, organic pollutants and P, in dissolved and particulate forms, are transported to receiving water bodies (Yang and Zhang 2011). The quality of surface runoff water is poor during flood events. The average concentrations of nitrate-nitrogen (NO₃^–-N), ammonium-nitrogen (NH₄⁺-N), total N, molybdate-reactive P, total P, and suspended matter in the runoff of Nanjing City, China were 8, 2, 6, 15, 6 and 4 times greater than those of the forested watershed, respectively. Barałkiewicz et al. (2014) observed that the Cd, Zn, Pb, and Co in the runoff from the largest impervious area had the strongest toxic effect on aquatic organisms. In the city areas, soil loses almost all of its inherent physical and chemical ability to filter, absorb and purify water. Therefore, a low infiltration rate due to soil compaction or sealing negatively affects the urban eco-environment, increases the rate of instantaneous floods and degrades surface water quality.

4.2. Heat exchange

Cities, on average, experience increased temperatures compared with surrounding rural hinterlands, which is known as the urban heat island (UHI) effect (Di Sabatino et al. 2009; Holderness et al. 2013). Cheon et al. (2014) found the increases in air and soil temperatures in rural areas were significantly lower than in the metropolitan cities in Korea between 1960 and 2010. The consequence of UHI is that cities will experience extreme temperature events with increased levels of mortality and human discomfort (Di Sabatino et al. 2009).

The UHI effects result from increased absorption and re-radiation of thermal energy from man-made materials, as well as anthropogenic heat outputs (Holderness et al. 2013). Anthropogenic heat is generated by human activities, besides the waste heat from air conditioning, industry and other sources, mainly as a result of replacing natural surfaces by buildings and pavements with different thermal inertia (Landsberg 1981). The sealed soil has lost its ability to adjust temperature. Weng et al. (2007) detected a higher surface temperature for impervious surfaces in the city of Indianapolis, USA, indicating that unsealed soils were essential for temperature regulation. Lazzarini et al. (2013) assessed land cover-temperature interactions in the Abu Dhabi metropolitan area over a 10-year period between 2000 and 2010 with a multi-sensor approach and confirmed this trend, adding the contribution of impervious surface areas to an average increment of 1 K during winter and 2 K during the summer. Xiao and Weng (2007) in China revealed that a change in land use toward urban impervious surfaces brought about an increase of air temperature in Guizhou province, China.

4.3. Gas exchange

In urban regions, beside the supply of water and nutrients, sufficient soil aeration is an important site factor for trees. The continuous air-filled soil pores are the only connection between roots and atmosphere. If soil pore volume and soil pore continuity are reduced by compaction or sealing, roots cannot be supplied with O₂ and, conversely, CO₂ emissions from the soil are inhibited (Herbauts et al. 1996; Horn et al. 2007). Until now, little information has been available on how the different types of sealant affect soil gas diffusivity or how urban trees react to soil aeration deficiencies. Against the background of existing investigations, soil aeration efficiencies hamper root respiration (Qi et al. 1994; McDowell et al. 1999; Gaertig et al. 2002), leading to the functional loss of fine roots (Rickman et al. 1965). Recent research in Kassel, Germany, showed that the lowest gas diffusivities and respiration rates were found at sealed sites, and the highest values were measured at vegetated sites such as lawn or flower beds (Weltecke and Gaertig 2012). Though soil gas diffusivity primarily controls soil respiration, soil CO₂ concentration is not strictly linked to the coverage type, and does not show a strictly directed dependence on topsoil gas diffusivity and soil respiration. The interesting result is that the investigated plants were not affected by soil aeration deficiencies (Weltecke and Gaertig 2012). Unexpectedly, Scalenghe and Marsan (2009) concluded that small portions of landscape sandwiched between sealed surfaces may contribute significantly to regional soil–atmosphere gas exchange, because unsealed patches within a urbanized region of the Great Plains, USA, occupied 6.4% of a 1578 km² area, but contributed up to 5% and 30% of soil methane (CH₄) consumption and nitrous oxide (N₂O) emission, respectively (Kaye et al. 2004).

Studies in functionally different urban areas showed the production of CO₂ by the soils of the industrial zone was lower by 1.3, 1.7, 1.8 and 1.8 times than that of the recreational or residential zones in Sergiev Posad, Shatura, Serpukhov and Serebryanye Prudy districts, respectively (Ivashchenko et al. 2014). However, it should be pointed out that soils of the urban ecosystems are capable of higher microbial respiration than the arable soils. Therefore, it can be supposed that the gas-producing activity in the urban areas will be comparable to that of the soils in the natural ecosystems (Ivashchenko et al. 2014). Soil compaction and sealing in the city areas seriously reduce or eliminate the soil gas capacity and exchange, while the uncompacted soil with vegetables will fully play its role for gas activities.

4.4. Human health

By characterizing the aerosol particle fingerprint based on nuclear microprobe, a study in Shanghai, China, showed that 31% of atmospheric aerosol was from surface soil (Qiu et al. 2001). Urban air dust is especially hazardous due to its carried pollutants and its direct impact on human inhalation. In developing countries, soil-borne dust has been a major source of air pollution with rapid city expansion. In industrialized areas, abandoned sites without plant cover can generate heavily contaminated dust and deteriorate air quality (Höke and Burghardt 2000). Up to now, little has been done about dust generation, transmission and its control in relation to urban soil conditions, or about the potential hazards related to soil-borne dust on human health.

The main concern about the biological effects of urban soil pollution is on human and ecosystem health. Urban soil pollution affects human health in two major pathways (Zhang et al. 2003b), i.e., the accumulation of pollutants in the urban agricultural soil–vegetable system and the successive transfer to the food chain, and direct exposure through touch and inhalation. Khillare et al. (2014) found that the incremental lifetime cancer risk via dermal contact and the ingestion pathway ranged between 10^–2 and 10^–4 while the risk due to inhalation pathways was between 10^–7 and 10^–9. Furthermore, they concluded that health risk for children was observed to be greater than that of adults, and was mainly contributed by ingestion and dermal routes. Hamad et al. (2014) estimated that the hazardous index based on inhalation and ingestion of potentially toxic elements was high for both children and adults. It is possible to estimate the total uptake of pollutants by a certain population, and to evaluate the potential health risks by a determination of vegetable consumption and pollutant content if supported by pollutant transfer models. Now, some exploring methods based on soil extraction and vegetable accumulation have been reported (Nabulo et al. 2012; Swartjes et al. 2013). Agbenin et al. (2009) found that leafy vegetables could accumulate metals in their tissues to unsafe levels even when total metal concentrations in these soils were below the allowable concentrations, in agricultural soils in northern Nigeria. Therefore, the systematic study of the transfer of soil pollutants through the food chain and direct inhalation uptake, and their impacts on human health, is very much needed and meaningful, as it can be used in (sub)urban soil management and agricultural product safety control.

Due to the difficulty in experimental and technical operation, little has been reported about the human health risk from direct dust inhalation. However, indirect studies such as children’s blood Pb testing showed that contaminated urban dust/surface soil was an important factor affecting human health in the urban environment (Calabrese et al. 1997; Jiang et al. 2004). Further studies showed that heavy metal contents varied by particle size, and fine soil particles were often more enriched with pollutants such as Pb, which could partly explain the increase in human blood Pb as fine particles could more easily become dust and be inhaled (Wang et al. 2006). In order to assess the bioavailability of urban soil pollutants, some studies developed extractants by simulating human/animal digestive fluid (Ruby et al. 1996; Ding and Hu 2014), which could better predict the uptake of pollutants by human/animal bodies (Oomen et al. 2003; Yamada et al. 2003). However, a more precise assessment of health risk by soil pollution needs to consider more than chemical indicators, and site features such as dust-producing conditions, dust density, range and inhalation amounts should be comprehensively included, but no such research has been reported yet.

5. SUMMARY

Urban soils are formed with significant anthropogenic influences. Physical, chemical and biological properties of urban soils are typically heterogeneous across various urban and suburban environments. With processes such as compaction, element enrichment and contamination, urban soils have direct ecological and environmental impacts. The lowering of urban soil quality is a sacrifice for the urban environment and the ecosystem as a whole. However, the physical and chemical degradation of urban soil may affect urban water and air quality as well as ecosystem health, instead of providing a service. Many studies have concluded that urban soil is often heavily contaminated, but the effects of such contamination on environmental quality and human health are far from clear. It is suggested that research on the environmental capacity and critical thresholds of urban soil pollution, pollutant migration and transfer dynamics, interactions between soil, water and air, and the biological effects of urban soil contamination is urgently needed, and is important for the establishment of a risk assessment system and for the better management of urban soil resources.

Acknowledgments

This study was supported by the Natural Science Foundation of China (grant nos. 40235054 and 41130530).