With migration from rural areas and the growth of towns and cities, more than 54 percent of the global population is now living in urban areas (FAO et al., 2018). As urban soils are subject to numerous and diverse anthropogenic activities, and may therefore have high levels of pollution, the assessment and monitoring of urban soils are of paramount importance. Green areas in cities provide multiple ecosystem services, from primary services such as the reservoir of nutrients and organic carbon in the soil, to food production, or the provision of cultural and recreational services (Bretzel et al., 2016). As green areas provide a potential exposure pathway for organisms and humans, soil pollution there requires particular attention. Meuser (2010), in his book Contaminated urban soils, presents an extensive review of the main contaminants in urban soils and their sources and is a recommended reference to have for more details on this issue.
Soil pollution in urban areas can derive from contaminants released from point or diffuse sources. As defined in the glossary, point-source pollution arises from discernible, confined and discrete source from which contaminants are discharged. Diffuse soil pollution arises from multiple points of origin and may be transferred long distances from other media. Major sources of point-source and diffuse pollution in urban soils are presented below or in the following sections of this chapter (see Section 3.5. on industrial and transport areas and 3.4.6Error! Reference source not found. on industrial and urban waste).
Houses and commercial areas contain hundreds of thousands of chemicals. From pharmaceuticals and personal care products (PPCPs), cleaning products, paints and solvents, flame-retardants, pesticides for in-house insects and gardening, the domestic environment is potentially exposed to a great variety of organic and inorganic contaminants. When discussing indoor pollution, consideration is given primarily to air pollution’s impact on human health, but rarely to the impact and the routes of entry into the soil and thus on the environment. However, even though the largest influx of contaminants into the soil from urban and commercial activities is through transport and waste, it is important to consider other entry routes.
One of the major concerns for human health and the environment is the presence of lead in soils, which in urban areas is mostly associated with the use of lead-based paints and transport emissions. Due to its high biotoxicity, bioaccumulation and persistence, no safe level of exposure to lead has been determined (WHO, 2019a).
Many developed countries banned the use of lead in fuel, paint and other products in the past century; for example lead in household paint was banned in Australia in 1920 (Needleman, 1998), while the United States of America banned it in 1978 (US EPA, 2013) and China in 1986 (Lin et al., 2009). Lead and its compounds are controlled in the European Union under the REACH regulation (Registration, Evaluation, Authorisation and Restriction of Chemicals), which came into force in 2007 (EU, 2015). The Eurasian continent has the highest proportion of countries that have included binding values of lead in paints and fuel in their national legislation, although it remains uneven in the region (Figure 8), with significant gaps in Eastern Europe and Central Asia (UNEP, 2019d). The international community has joined forces to eliminate lead from all paints by 2020 through the Global Alliance to Eliminate Lead Paint headed by UN Environment and WHO (UNEP and WHO, 2011). According to their data, in 2019, 73 countries around the world already had a regulation on acceptable levels of lead in paint (UNEP, 2019d). Since 2009 IPEN, a member of the Global Alliance, has analysed more than 100 studies from 59 countries and found that lead paints are still widely sold in low- and middle-income countries. The studies comprised more than 3 300 solvent-based paints. Many of these paints contained very high levels of lead above 10 000 parts per million (ppm) of the dry weight of the paint. The results can be seem on IPEN’s interactive map (IPEN, 2020).
However, despite national and international regulations, old houses and even newly produced paints still have high concentrations of lead and are widely commercialized mainly in low- and middle-income countries (LMIC) (Clark et al., 2006; IPEN, 2017; Lin et al., 2009; O’Connor et al., 2018). Lead from these sources ends up in yards and gardens as the paint ages or after being intentionally removed (Nezat, Hatch and Uecker, 2017; Shayler, McBride and Harrison, 2009), with the accumulation of lead in soils to levels exceeding the permitted values (Mielke et al., 2004; US EPA, 1996). Other sources of lead pollution in urban environments include lead piping and tanks in old domestic plumbing systems (Beattie et al., 1972), vehicle exhausts in countries that still allow the use of leaded fuels (see section 3.5.5), and deposition of airborne particles from lead emitting industries (see section 3.5.2). The elimination of lead from petrol in many countries has led to a reduction of 39 percent of blood lead levels worldwide (JECFA and World Health Organization, 2011).
In natural areas, where hunting is practised, the presence of lead in soils released from the weathering of lead-containing ammunition is also a concern for human health and the environment (Duggan and Dhawan, 2007; ECHA, 2018).
Despite lead being the most concerning trace element in urban areas due to its high toxicity, other trace elements may also be found in urban soils. Zinc and copper are mainly associated with atmospheric deposition of transport emissions (see Section 3.4.5), including vehicle exhaust gases, tyre and brake abrasion residues, and road wear emissions (Ferreira et al., 2016; Kadi, 2009; Lee et al., 2006), while mercury is frequently linked to coal combustion (Luo et al., 2015; Pan et al., 2016) and industrial emissions (see Section 3.5).
Phthalate esters/phthalic acid esters (PAEs) are used as plasticizers and additives in many industrial products, PPCPs and food containers. PAEs have been found to be ubiquitous in all soils as they can have high volatility (Wu et al., 2015b), which raises a serious concern, as their role as endocrine disrupters has been widely documented (Casals-Casas, Feige and Desvergne, 2008; Ghisari and Bonefeld-Jorgensen, 2009). A high concentration of low molecular weight PAEs (dimethyl phthalate) was observed in residential soils in Xi’an city, in China, associated with the use of PPCPs (Wang et al., 2018); while in urban soils of Paris, France, diethyl phthalate, di-n-butyl phthalate, and di-iso-nonyl phthalate were the predominant PAEs, reaching values 12, 10 and 7 fold higher than those observed in agricultural areas, respectively (Tran et al., 2015). The PAE content of Moscow’s soil also supports the hypothesis that urban emissions are major sources of PAEs, only being surpassed by soils where a landfill exists (Brodskiy et al., 2019).
Brominated flame retardants (BFRs) are frequently used in electronic circuits, building materials and textiles and hence can be found in urban soils (Li et al., 2018b), especially at waste and industrial sites (Tang et al., 2014). This has resulted in the pollution of eggs and cows’ milk produced around dump sites (Oloruntoba et al., 2019). As mentioned in Rahman et al. (2001), BFRs are found in high-impact polystyrene, flexible polyurethane foam, textile coatings (not clothing), wire and cable insulation, electrical and electronic connectors and other interior parts. Its production and use was intensive during the last decades of the last century, but many of the industrial products containing BFRs may still in use or have been recently discarded (Stockholm Convention, 2017b). Polybrominated diphenyl ethers (PBDEs) are the most commonly used BFRs. They are persistent lipophilic compounds with a similar structure to polychlorinated biphenyls (PCBs) that can bioaccumulate and have been reported as endocrine disruptors (Patisaul et al., 2013; Skarman et al., 2005). Due to potential toxicity, PBDEs were included in the Stockholm Convention List (Stockholm Convention, 2017c), therefore all parties must eliminate their production. However, some exemptions have been adopted and PBDEs are still in use (Stockholm Convention, 2017a).
Urban and peri-urban agriculture is growing rapidly due to the increase in urban population and food demands. According to FAO, some 800 million people worldwide produce food in urban gardens (FAO, 2019a), which has a major beneficial impact on household incomes, food security and nutrition while providing urban green space and its resilience to climate change. However where the soil is polluted, special attention must be paid to the potential transfer of contaminants from urban soils to crops and the risk posed to human health.
POPs (including organochlorine pesticides (OCPs), PCBs, PCDD/Fs, PBDEs, and PAHs), and trace elements may be found in urban horticultural gardens and urban parks (Jiang et al., 2011; McGrath et al., 2016). They pose a potential risk to human health due to bio-accumulation of these contaminants in cattle and chicken (Augustsson et al., 2015; Weber et al., 2019, 2018a), and continued contact with contaminated soils during recreational activities, which is especially dangerous for young children. (Augustsson et al., 2015) found elevated concentrations of trace elements in lettuce and potatoes grown in home gardens in southern Sweden, in an area heavily polluted due to former glass manufacturing industry, resulting in daily intakes of cadmium exceeding the WHO maximum tolerable intake values (WHO and FAO, 2011). Similar results were obtained by Mombo et al. (2016), who observed cadmium and lead uptake in lettuce in home gardens in France, although with a limited transference to humans. Concentration of trace elements in soil alone cannot predict plant uptake and the potential risk to human health, as there are other factors involved, such as soil properties (see Chapter 2) or plant characteristics and metabolism. However, it is necessary to take into account the total and bio-available fraction of trace elements in soils when performing risk assessments. It is also recommended to monitor the population consuming produce from urban gardens in contaminated soils, since direct exposure to contaminated soil particles together with the intake of food with high trace elements contents can have a cumulative effect over time.
Synthetic fertilizers and pesticides for home and small-scale use are available at local stores. The types and quality of products available for purchase will depend on national chemicals legislation and its enforcement, and the prevalence of illegal and counterfeit products (Haggblade et al., 2019). Cheng et al. (2015) reported high levels of lead, chromium, arsenic, and cadmium in soils of community and private gardens in New York city, all of which represented a high risk to human health (Paltseva and Cheng, 2019). Lack of knowledge of the most appropriate rates and application methods often leads to excessive application of agrochemicals that end up accumulating in soils. Urban horticulture plots are subject to sources of pollution similar to those in agricultural areas (see Section 3.2), such as excessive application of agrochemicals, irrigation with low quality or recycled water and the contribution of home-made compost rich in contaminants (Tixier and de Bon, 2006).
Pesticides are also widely used to control weeds and pests that affect vegetation in urban green areas or to control disease vectors. Under Annex B, the Stockholm Convention allows countries to apply for an “acceptable purpose” exemption to use DDT to control of the malaria-transmitting Anopheles mosquito generally by indoor residual spraying (Stockholm Convention, 2019). Therefore, high levels of highly toxic and persistent organic contaminants have been found in public parks around the world (Table 4).
Worrying examples of high OCPs content in urban areas have been reported in different parts of the world, whether from previous or current use. In Beijing’s urban parks, Li et al. (2008) found that total DDT levels in Zhongshan Park exceeded the Chinese environmental quality standard for soils (Ministry of Ecology and Environment, 2018). Also in urban parks of Moscow, total DDT levels ranged from 3.78 to 1 347 μg/kg (mean value 143.1 μg/kg) (Brodskiy et al., 2016). In commercial areas of Kathmandu, high levels of DDT have been observed, even higher than those detected in transport routes (Pokhrel et al., 2018). Practically all soils of Warsaw city considered in the study carried out by Bojakowska, Tomassi-Morawiec and Markowski, (2018) showed high contents of DDT, even in urban horticulture plots, with concentrations that exceeded, in some cases, 5 000 μg/kg of total DDT. DDT is the most regulated pesticide in the world, with threshold values for soils that range from 0.33 μg/kg in Oregon to 10 μg/kg in the Netherlands (Li and Jennings, 2017). The predominance of p-p’ DDT in soils has been observed in many countries, including Mexico (Wong, Alegria and Bidleman, 2010), Nepal (Pokhrel et al., 2018), Kenya (Sun et al., 2016) and India (Mishra, Sharma and Kumar, 2012), indicating that DDT is still being used and it is commercially accessible in many parts of the world.
Accidents and incidental leakages constitute another important source of soil pollution in urban areas. Fuel tanks and pipelines, sewers, septic tanks and landfills are large reservoirs of a wide variety of contaminants. Leakage from power transmission equipment can contain PCBs. They all can suffer accidental or incidental losses due to natural disasters, such as the action of hurricanes, earthquakes or fires, due to poor maintenance causing deterioration and consequent rupture and leakage, as well as external damage during excavation (Ramírez-Camacho et al., 2017).
The presence of PAHs in urban soils is highly related to oil spills although they may also derive from combustion emissions (Wang et al., 2017a). Wang et al. (2013) in their study of PAHs in urban soil in Shanghai identified six sources of the total soil PAH burden: petrogenic sources (6 percent); coal combustion (21 percent); biomass burning (13 percent); creosotes (16 percent); coke production (23 percent) and vehicular emissions (21 percent). The relationship between oil spills and PAHs in the environment is clearly demonstrated by their presence near research stations in Antarctica, land that would otherwise remain virtually uncontaminated (Aislabie et al., 1999). According to Ivshina et al. (2015), the prevention of oil spills should be a priority worldwide, including the strengthening of cooperation between neighbouring countries, due to their high frequency and potential effect on human health and the environment. The presence of oil in soils seriously affects plant growth and metabolism (Adam and Duncan, 2002). Han et al. (2016) observed a reduction of gas exchange and photosynthesis rate in legumes growing in oil-polluted soil, while Tran et al. (2018) observed significant negative impact on germination rate of two acacia species. Changes in soil microbial communities has been also reported in oil-polluted soils, where oil degrading bacteria tend to predominate over native species (Abbasian et al., 2016; Jia et al., 2017).
Leakage from underground fuel tanks at vehicle filling stations is also very common, with the plume of contaminants spreading through soil with the groundwater. In 1984, the U.S. government set up a programme to address these leaking underground storage tanks. By 2010, 588 000 leaking underground storage tanks had been identified of which 80 percent had been remediated (Congress Research Service, 2010).
As discussed in Section 3.3.4, sewers lines and septic tanks are associated with domestic and hospital wastewater discharges and are therefore rich in PPCPs, faecal bacteria and other pathogens, as well as antimicrobial-resistant bacteria, micro and nanoplastics and poly- and perfluoroalkyl substances apart from more conventional contaminants such as POPs or trace elements (Ellis, 2006). A leakage from wastewater systems will release a variety of contaminants into soils that may ultimately reach groundwater (Rutsch et al., 2008; Wolf, Zwiener and Zemann, 2012). The fate and risk of emerging contaminants such as PPCPs to human health and the environment are only starting to be elucidated. In addition, research of contaminants in wastewater has mainly focussed on monitoring contaminants in ground and surface water bodies and little attention has been put to their retention in the soil matrix (Foolad, Ong and Hu, 2015).
Contaminants in urban soils may also have a more distant origin (diffuse pollution) and come from atmospheric deposition of emissions from industrial and artisanal activities (Douay et al., 2007), municipal solid waste incineration (MSWI) (Ma et al., 2018b; Weber et al., 2018a), or mining activities (Liang et al., 2017) from areas surrounding urban centres. Where pollution from atmospheric deposition falls in catchment areas, it can be transported and concentrated by runoff and floods. The distance from such air discharge activities, the predominant wind direction and the weight and size of the particles, are all influences on the soil contaminant content (Li et al., 2019; Meuser, 2010). This is illustrated in the Janghang Smelter case study in Chapter 7. Contaminants such as inorganic mercury (IOMC and IOMC, 2002; Pacyna et al., 2016; Pearson et al., 2019; UN Environment and AMAP, 2019), POPS and other organochlorine compounds (Balmer et al., 2019; Cabrerizo et al., 2018a) are capable of long-range atmospheric transportation before deposition (Figure 9). Atmospheric transport of contaminants occurs transnationally along all cardinal axes, north-south and east-west (Cabrerizo et al., 2018b; Hung et al., 2016).
With respect to industrial and mining activities Song et al. (2018) and Jamshidi et al. (2007) observed a clear pattern of decreasing PCB in soils as the distance from the source increases in China and the United Kingdom of Great Britain and Northern Ireland. Similarly, an inverse correlation was observed between the content of trace elements in soils, mainly cadmium and lead, with their distance from a former smelter in France (Douay et al. 2007), indicating a limited but significant mobilization of trace elements in the atmosphere. In a similar study, an irregular pattern of trace element distribution was observed in urban areas of Chicago that Cannon and Horton (2009) associated with atmospheric deposition from different anthropogenic sources, the impact of which was significantly greater than in surrounding agricultural areas. The analysis of trace element sources and distribution pattern at a distance of 3 km around a municipal solid waste incinerator in North China also supports the hypothesis of atmospheric transportation and diffuse pollution (Ma et al., 2018a). Atmospheric deposition must therefore be taken into account when conducting risk assessments in all soils (rural, urban and industrial), where there is reasonable expectation that there could be significant deposition of contaminants.
Erosion runoff and flooding also represent a major source of diffuse soil contamination and pollution to and from urban soils and neighbouring soils under other land uses (Boardman et al., 2019). Organic contaminants and trace elements are strongly bound to soil organic matter and clay minerals, and hence can be easily transported downstream (Rodríguez Eugenio, McLaughlin and Pennock, 2018). Reducing (anaerobic) conditions in soils during flooding and waterlogging can lead to the mobilization of trace elements as a secondary effect. Such conditions release iron and manganese from soil oxides and organometallic complexes (Mukwaturi and Lin, 2015). An example of mobilization of contaminants and downstream migration is presented by Dias-Ferreira et al. (2016), in which high levels of trace elements and PCBs were detected in alluvial soil profiles. Paule-Mercado et al. (2016) found downstream transport of faecal bacteria in urban areas, whose origin was associated with pet faeces and leachate from sewers. Salinization of downstream soils is also associated with urban pollution; Robinson and Hasenmueller (2017) observed the temporary salinization of soils following runoff events and flooding after the application of salt on roads to prevent icing.
Global population has shifted from a predominantly rural to an urban majority (Figure 10), and it is in urban areas where the generation of municipal solid waste (MSW) associated with lifestyle and access to packaged products is concentrated.
Official statistics and definitions of MSW are diverse globally, and in many countries there are no standardized reporting mechanisms for municipal waste generation and its management. This makes it difficult to estimate the amount of waste produced at the global level and its polluting potential (Kawai and Tasaki, 2016; UNECE and Statistics Netherlands, 2017). Kaza et al. (2018) estimated the global production of MSW at 2.01 billion tonnes per year, but it may be much higher. Additionally, unconventional waste collection and recycling is often not included in any statistics (Mian et al., 2017), despite being important in some countries and even being part of the local economy. Municipal waste management has become one of the humanity’s most challenging environmental problems (UN-Habitat, 2010).
Municipal solid waste includes household waste such as packaging, food scraps, grass cuttings, furniture, clothes, paper and cardboard, household appliances, paint and batteries. It may also include similar waste from small commercial activities, office buildings, educational facilities and governmental buildings, and small businesses that treat or dispose of the waste in the same facilities used for waste collected by the municipality. However, the definition greatly varies among countries even in regions under the same legal framework such as the European Union (Kawai and Tasaki, 2016). While some countries consider e-waste, medical or hazardous waste within the municipal solid waste stream, this report addresses them separately.
According to Kaza et al. (2018), globally, MSW is on average composed of 44 percent food and vegetable waste, 17 percent paper and cardboard and 12 percent plastics and, to a lesser extent, glass, metals, rubber and leather, and wood (Figure 11), although there are considerable differences depending on the country’s income level. A large part of this waste is not hazardous to the environment and the health of workers and neighbouring populations if it is managed in an environmentally sound manner. However, due to differences in MSW definition or less strict regulations and controls, hazardous household waste and industrial waste from small enterprises could enter the MSW system (Azeez, Hassan and Egunjobi, 2011; Miezah et al., 2015). Some hazardous wastes will enter the MSW stream even if appropriate mechanisms are in place for their separate collection and disposal, either because the population is unaware of the existence of these mechanisms or because of a lack of awareness of the damage they can cause to the environment (Ferronato et al., 2017). This kind of co-disposal of hazardous and non-hazardous waste is now regulated in many developed countries, such as the United States of America (US EPA, 2016b), the European Union (EC, 2019b), and Japan (MoE, 2014), but still continues in other regions. Lemanowicz, Bartkowiak and Breza-Boruta (2016) detected a high impact of mixed waste (hazardous and non-hazardous) on microbial activity and soil properties in illegal dumping in Poland.
High-income countries are attempting to treat MSW as a resource and to increase the value that can be recovered from it (Figure 12). Most European countries require households to segregate their wastes into recyclable fractions, organic fraction for composting, hazardous fractions for specialized disposal, and the un-recyclable remainder, which often is converted to refuse-derived fuel for power generation. These approaches are driven by economics, international environmental commitments, concern over the sustainability of landfill as a disposal practice and the opportunity costs of the land that it occupies. Many countries do not have sufficient infrastructure to recycle and treat their own wastes, and rely on exporting them for treatment in other countries. This has caused issues for some of the receiving countries whose waste recycling and treatment capacity became overloaded. China banned the import of most plastic waste at the beginning of 2018, which caused stockpiles to accumulate in US, Europe and Australia, and the migration of the recycling trade to less developed countries (Parker, 2018). In 2019 parties to the Basel Convention agreed that plastic waste should be included in its legally binding framework in order to make the global trade in plastic waste more transparent and better regulated, whilst also ensuring that its management is safer for human health and the environment (Basel Convention Secretariat, 2019).
Although landfill is one of the least preferred disposal options on the waste management hierarchy (Figure 13) (Hyman et al., 2015), it remains a common practice throughout the world. Landfill sites can vary significantly depending on their design and management. In their report, Kaza et al. (2018) used the three classifications of landfill: sanitary landfill, controlled landfill and uncategorised landfill. Sanitary landfills have an impermeable base, leachate capture and treatment, and landfill gas capture and treatment. Controlled landfills are engineered but do not have landfill gas collection systems. Unclassified landfills are those where there was no information about the leachate or landfill gas collection systems. They used the term open dump to describe the uncontrolled dumping of waste on land, in waterways and marine environments.
Kaza et al. (2018) reported that 70 percent of MSW was disposed of in landfills, whether in controlled landfills or sanitary landfills, or in open dumps as is shown in Figure 12. The prevalence of the use of landfills as a disposal option and the practices to manage them varies significantly among countries and regions dependent on their income levels. In North America, landfill is only used for 54 percent of MSW, and they are all sanitary landfills, which should have effective leachate and landfill gas management systems. While in a region with more developing countries such as South Asia, 79 percent of MSW is landfilled, of which the majority are classified as open dumps. Open dumps tend not to have effective leachate and gaseous emission management and many are subject to fires, with the release of toxic emissions (Renou et al., 2008). In Low and Middle-Income Countries (LMIC), uncontrolled dumping and open burning of waste in cities, around houses, in brownfields, and other areas outside cities are common practices (Medina, 2010; Mmereki, 2018), which have been also reported in upper middle-income countries, as it is the case of China (Mian et al., 2017), Albania (EEA, 2018), or South Africa (Rasmeni and Madyira, 2019).
Poor design and management of MSW landfills increases the potential for pollution of soil through leachate and emissions to atmosphere (Renou et al., 2008). In these uncontrolled MSW dump sites, the presence of a cocktail of contaminants in the leachate poses a risk to the environment (air, fresh water, soil, and biota) and human health (Barčić and Ivančić, 2010; Devkota and Watanabe, 2005; Ferronato and Torretta, 2019; Njagi et al., 2016; Owusu Boadi and Kuitunen, 2002; Samadder et al., 2017; Talyan et al., 2007; Wangyao et al., 2010; Windom, 1992). Leachate characteristics depend on the composition of the solid waste, the physicochemical properties of the waste pile, including hydrology, the age of the landfill, and the available oxygen, as well as climatic conditions (Yusoff et al., 2013).
Leachates from young landfills are rich in dissolved organic carbon (DOC) (Reinhard, Barker and Goodman, 1984), which causes soil and groundwater pollution, since this DOC is capable of leaking from the waste pile and mobilising some organic contaminants and trace elements contained in the landfill. Concerning amounts of trace elements such as cadmium, lead, manganese and zinc have been observed in soils beneath landfill sites in Nigeria and Pakistan (Aiman et al., 2016; Olayinka et al., 2017). Denton et al. (2016) have also reported high levels of cadmium, copper, lead, mercury and zinc in co-disposal landfills in Saipan (Mariana Islands). However, it has been observed that degradation of organic matter stabilizes after some years, depending on aeration and climatic conditions, and contaminants are retained within the stable organic fraction and therefore are unlikely to be mineralised or mobilised in the long-term if waste is rich in organic matter (Chen, Knappe and Barlaz, 2013; Keuter, Fründ and Kerth, 2011; Lemanowicz, Bartkowiak and Breza-Boruta, 2016).
For example, in the Russian Federation, where about 97 percent of MSW is disposed of in landfills, mixtures of trace elements, organic contaminants, such as PCBs, and pathogens were found in soils adjacent to landfills (Zamotaev et al., 2018). The proliferation of pathogenic organisms, such as viruses, faecal bacteria, protozoa and helminths, which can cause infectious diseases to soil biota and humans, has been reported in many unmanaged landfills (Gerba et al., 2011; Kalwasińska and Burkowska, 2013 and references therein). These authors also referred to the potential for airborne transport of pathogens in the vicinity of landfill sites. The majority of these pathogenic agents came from pet faeces.
In addition, MSW is considered a source of emerging contaminants, including antimicrobials, from PPCP, baby diapers and toilet paper, or PFAS (mainly PFOS and PFOA) present in cleaning products, textiles, carpet, treated paper and plastic food packaging, cosmetics, and fire retardants, among others (Eggen, Moeder and Arukwe, 2010). All these emerging contaminants potentially pose a serious risk to the environment and human health even at trace concentrations, their identification and control is therefore of primary importance. Antimicrobial resistant (AMR) bacteria and genes have been widely reported in leachates from MSW (Threedeach et al., 2012; Wu et al., 2017; Yu et al., 2016; Zhang et al., 2016b). Considering that unmanaged dumping sites and poorly controlled landfills are mostly located in developing countries, where control measures for the sale and disposal of antimicrobials are much less stringent, it is possible to conclude that it is in these countries that there is the greatest risk of transfer of AMR genes and therefore the greatest impact on human health (Wu et al., 2017).
Slack, Gronow and Voulvoulis (2005) reviewed the organic contaminants in landfill leachate in different parts of the world and their sources and found them to include phthalates, chlorinated aromatics, bisphenol A, analgesics and antimicrobials, detergent surfactant, and trace elements. Also a major share of wastes containing PFOS, PFOA and other PFAS have been disposed in landfills (Weber et al., 2011) and are released in landfill leachates (Busch et al., 2010; Lang et al., 2017; Li et al., 2012a). The authors concluded that household waste often includes products that may be considered hazardous because of the residual levels of contaminants, such as lead-based paints, garden pesticides, pharmaceuticals and personal care products, cleaning chemicals and detergents, fluorescent tubes, used oils, batteries and accumulators, highly flammable aerosols, treated wood, small appliances and electronic equipment.
But it is not only leachate that poses a risk to the environment and human health (Figure 14). Open burning of household waste and MSW is a common practice worldwide (Zhang et al., 2011a), as a strategy to reduce the amount of waste to be managed or transported to landfills. Waste incineration is also gaining popularity in many countries, mainly in high income countries but also in middle income countries, accounting for 11 percent of waste management globally (Kaza et al., 2018). PCDD/Fs, PCBs and PAHs are often found in areas surrounding the open burning of municipal solid waste and incinerators not operated with best available technology (BAT) (Cheruiyot et al., 2017; Oh et al., 2006a; Weber et al., 2018a), as identified in the inventories of the Stockholm Convention (Fiedler, 2007, 2016). The nature of waste and the presence of contaminants, such as pesticides or flame retardants, will determine the release of these organic contaminants during and after burning (Gullett et al., 2010; Zhang et al., 2011a, 2011b). Chakraborty et al. (2016) identified the sources of PCBs in Indian soils, of which 35 percent was associated with municipal and medical solid waste incineration and open burning of waste.
Additionally, informal waste collection and recycling represents an income for the poorest families that gather around the large cities and engage in the extraction of valuable waste from open-air dumps or the informal recycling of useful minerals. The majority of materials recovered by the informal sector include plastics, metals and cardboard. In Mumbai city, for example, about 150 000 waste pickers collect practically all plastics produced for recycling in informal recycling industries and workshops (Vaidya, Kumar and Sharma, 2016). Therefore, this informal recycling sector contributes to reduce the amount of waste in landfills and generate an added value over used items (Agarwal et al., 2005; Tarekegn, Fantahun and Tefera, 2016). However, this informal recycling and processing can also have an adverse effect on the environment and waste pickers’ health since it is frequently done without any kind of control or containment measure for the release of contaminants, which accumulate in the soil of the work areas, which are often the same ones in which they live, putting at serious risk not only the health of the recyclers but also of their families, and especially their children (Auler, Nakashima and Cuman, 2014; Coelho et al., 2016; Gutberlet and Baeder, 2008; Parveen and Faisal, 2005) (see Chapter 4 for more details).
According to WHO the typical composition of health-care waste (HCW) is 85 percent general non-hazardous wastes, 10 percent infectious waste and 5 percent with chemical and radioactive waste (WHO, 2014, 2017a). The hazardous fraction is described in (Table 5). Although not specified in the WHO definition, similar wastes from veterinary centres should also be considered in this category.
Global data regarding the production and management of health-care waste is absent or very incomplete and only scattered data can be obtained for single countries (Caniato, Tudor and Vaccari, 2015; WHO and UNICEF, 2015). WHO (2019) estimates that in high-income countries the average daily generation rate of hazardous waste is 0.5 kg per hospital bed; while in low-income countries it is 0.2 kg. However much higher daily generation rates have been reported, with 2.8 kg/bed in Ghana (Adama et al., 2016), between 0.5 and 2 kg/bed in India (Mathur et al., 2011) and 2.5 to 3.01 kg/bed in the Islamic Republic of Iran (Farzadkia et al., 2009). With a growing population whose life expectancy has increased considerably over the last few decades, the demand for medical services and the generation of HCW will tend to increase in the coming decades (Olaniyi, Ogola and Tshitangano, 2019).
Due to their hazardous nature, the environmentally sound management of HCWs is essential to reduce the risk to human health and the environment. WHO has developed guidelines for the safe management of HCW that has been updated and improved over the years (WHO, 1999, 2014, 2017a). The Basel Convention on the control of transboundary movements of hazardous wastes and their disposal includes references, guidelines and obligations to the member countries about environmentally sound management of HCW (Basel Convention, 2016). The Stockholm Convention also includes guidelines and recommendations to reduce the production of POPs from waste incinerators (Stockholm Convention, 2012). More recently, the Minamata Convention on Mercury provides for the reduction of certain medical equipment containing mercury, such as thermometers and blood pressure measuring devices (UNEP, 2019b; WHO, 2015).
The relevance of HCW management to human health is evident and has been taken into account in all these international approaches, while its impact on the environment, in particular the soil, has often been overlooked. This is clearly reflected in the lack of consideration of the soil (land) as a receptor of the organic contaminants emitted after the incineration of HCW (Stockholm Convention, 2012), despite the evidence of the impact of fly and bottom ashes on soils (Sabiha-Javied, Tufail and Khalid, 2008; Tanjim Shams et al., 2012).
In the event that WHO guidance (WHO, 2017a) on the safe management of hazardous HCW is not followed there is a risk that soil could become polluted. Hazardous waste that is dumped or disposed of in a poorly designed and managed landfill risks pollution of surrounding soils through migration of hazardous contaminants through leachate and physical dispersal by wind, scavengers, and human activity.
Open burning or burning in a poorly designed and managed incinerator risks the dispersal of contaminants including fly ash and gaseous emissions that will deposit on the surrounding soil. Incomplete combustion of incinerated waste can release POPs and other organic contaminants that are capable of polluting soils at great distances through long-range atmospheric transfer (Allende et al., 2016; Yan et al., 2011).
The risks of soil pollution from the inappropriate disposal or recycling of general non-hazardous HCW are the same as those for municipal wastes and wastewater that are elaborated in Sections 3.3.6 and 3.2.4, respectively.
High-income countries reduce the risk from HCW by using treatment technologies such as autoclaves, microwaves, steam treatment or incinerators equipped with filters to prevent harmful emissions. However, the acquisition and maintenance of these technologies are not affordable for many health centres and hospitals in LIMCs (WHO, 2019b), which often lack the technical capacity to separate hazardous waste and manage it appropriately. Therefore, HCW is usually managed off site by private companies (Jang et al., 2006; Olaniyi, Ogola and Tshitangano, 2019). In some countries the sound management of HCW is hindered by a lack of awareness among medical staff of the risks that it poses to public health and the environment (ISWA, 2007; Khan et al., 2019; Patwary, O’Hare and Sarker, 2011; Sahiledengle, 2019; Sharma et al., 2013).
In LMICs, HCW is frequently deposited in open dumps and unmanaged landfills, which spreads the hazards to other non-hazardous wastes and poses a risk to scavengers, children and the environment as reported by Lekwot et al. (2012) in the Kaduna State, Nigeria. Burning HCW is another common practice in LMICs (Adama et al., 2016; Ali et al., 2017; Yazie, Tebeje and Chufa, 2019; Zhao et al., 2010). It can be useful to reduce waste infectiousness and volume. However, contaminants such as PCDD/Fs, PCBs and PAHs can be released during the process if it is done in the open or in an unsuitable waste incinerator (Jang et al., 2006; Singh and Prakash, 2007). Trace elements are concentrated in the ash, which, if disposed of inappropriately, can contaminate soil (Sabiha-Javied, Tufail and Khalid, 2008; Zhao et al., 2010). Adama et al. (2016) found significant levels of pollution by trace elements in soils within 60 m of an incinerator and neighbouring landfill in Ghana, while higher levels of PCDD/Fs were found in soils within 200 m of an HCW incinerator after three years of operation in China (Li et al., 2012b). Nevertheless, because these types of incinerators are generally located in industrial areas with heavy traffic or near landfills, identifying specific sources of trace elements and organic contaminants is not an easy task due to the similarities in compound profiles (Auta and Morenikeji, 2013; Ferré-Huguet et al., 2007; Jimenez et al., 1996).
Medical imaging equipment for diagnosis, research and radiotherapy treatments use different radionuclides with short and long half-lives (Qaim and Spahn, 2018). Short-lived radionuclides can be treated as non-hazardous waste after short-term specialized storage while decomposition occurs (Laraia, 2015; Ravichandran, 2017a). Long-lived radionuclides, such as 131I, a highly radiotoxic isotope used in nuclear medicine (Sundell-Bergman et al., 2008), require a long-term management strategy to avoid its impact on biological systems (Evdokimoff et al., 1994). Both hospitals and patients produce radioactive hazardous waste (Evdokimoff et al., 1994; Ravichandran, 2017b; Sundell-Bergman et al., 2008) and specific segregation and containment measures have to be developed for radioactive waste treatment in situ. While developed countries have concrete legislation and technologies to assess radioactive potential of medical waste prior their acceptance in waste management facilities, developing countries still need to improve their legal framework, technical capacities and awareness to reduce the impact of this hazardous waste (Al-Khatib, Al-Qaroot and Ali-Shtayeh, 2009; Dzekashu et al., 2016; Mollah et al., 2016).
Electronic waste or e-waste includes all electrical and electronic devices, including computers, mobile phones or household appliances (Townsend, 2011). As speed of development of high-technology products increases and they become more affordable, their working life shortens as they are replaced more frequently. The consequence is a rapid increase in the production of e-waste. According to Baldé et al. (2017), e-waste production accounted for 44.7 million metric tons in 2016 and it is expected to increase by about 14 percent by 2021.
E-waste contains valuable and rare minerals, such as gold, copper, neodymium or indium but also trace elements such as cadmium, chromium, lead and mercury, and persistent and emerging contaminants such as POPs, PCBs, plastics, perfluorooctanoic acids (PFOAs), and flame-retardants, among others (Realff, Raymond and Ammons, 2004; Hashmi and Varma, 2019).
Many high-income countries have legislation, often based on extended producer responsibility that aims increase the rate of collection and environmentally sound management of e-waste. However for economic reasons, much of the e-waste was exported to developing countries. It was estimated that only about 20 percent of e-waste was recovered and recycled properly globally (Baldé et al., 2017), thus 80 percent of e-waste ended up in urban landfills or dumpsites (Figure 15), which are mainly located in China, India, Mexico, Brazil, Eastern Europe and Central Asia, and West Africa (Lundgren, 2012). Receiving countries usually have less regulation on e-waste management and waste is frequently accumulated in unmanaged makeshift dumps. Many of the urban poor who make their living by reassembling obsolete devices for reuse or recycling them to extract valuable minerals and materials are exposed to toxic contaminants present in waste and soil, especially affecting women and children (Heacock et al., 2016; ILO, 2014). Although Baldé et al. (2017) does not include informal recycling and reassembly in the statistics of e-waste recycling, it is a considerable activity in receiving countries and is one of the main sources of soil contaminants associated with waste streams in these countries (Davis and Garb, 2015; Yang et al., 2018). Good progress towards reducing pollution in recipient countries has been made following the entry into force of the Ban Amendment of the Basel Convention on 5th December 2019, which prohibits developed countries to export hazardous waste to developing countries (Basel Convention, 2019).
Many studies have been carried out on the presence of contaminants in soils around uncontrolled e-waste recycling and processing facilities in China. Until the entry into force of the ban, China was the greatest importer of e-waste. From its growing domestic consumption, it also is expected to become the largest producer (Fu et al., 2018). Sepúlveda et al. (2010) conducted a review of heavily polluted areas from e-waste recycling facilities in China and India. In an e-waste open-burning facility in southern China, extremely high levels of brominated flame retardants, PAHs, PCBs and trace elements were measured in soils (Wong et al., 2007). However, the main concentrations were localized within 500 m from the e-waste open-burning site, except in the case of volatile compounds (PAHs). The high accumulation of brominated flame retardants in the topsoil in a local dismantling workshop in Qingyuan County and the secondary transfer in chickens due to their feeding habits was reported by Luo et al. (2009). Table 6 presents trace element concentrations measured in soils at various informal e-waste recycling facilities around the world.