HUMAN EXPOSURE TO CONTAMINANTS IN SOIL

Contaminant Fate and Transport in Soil

Once chemicals enter the soil, their movements and potential impacts are determined by a range of physicochemical characteristics, such as solubility, volatility, and sorption potential, as well as their interaction with specific soil properties. These interactions and processes (Figure 6-2) govern whether contaminants will migrate deeper into the soil through percolation, escape into the atmosphere via volatilization, or remain bound to soil particles (Jury and Godhrati 1989; Linn et al. 1993; Schnoor 1996).

FIGURE 6-2

The route of entrance and fate of contaminants in soils, and nine major soil processes determining the fate of soil contaminants. SOURCE: Food and Agriculture Organization of the United Nations. Reproduced with permission. Adapted from Current Opinion (more...)

Transformation of contaminants—loss or removal of certain substances (e.g., reactants) and the generation or formation of new substances (e.g., products)—can occur via both abiotic and biotic processes (Linn et al. 1993; Scow and Johnson 1996; Al-Mamun 2017). Physical and chemical transformation processes include photolysis and hydrolysis reactions (Figure 6-3). Coupling or polymerization reactions, mediated by certain mineral oxides and fungal extracellular enzymes, can also transform and alter the availability of some organic contaminants (Linn et al. 1993). Biodegradation specifically refers to the transformation of a contaminant by metabolic reactions carried out by the soil organisms, primarily the microbiome. Given that the soil microbiome, with its strong connection to soil health, is a major focus of this report, the committee chose to focus on biodegradation and its role in contaminant removal.

FIGURE 6-3

Transformation processes affecting the fate of pesticides and their degradation products in the environment. NOTE: In this example, a pesticide stands in as an example of an organic contaminant. SOURCE: Adaptation used with permission of Spring Nature (more...)

Human Exposure to Contaminants in Soil

The level of threat that a contaminant poses to human health is thus determined by the complex interplay between the contaminant and the unique soil conditions at a particular location. In general, the enhancement of soil quality (for instance, increasing the level of soil organic matter) can serve as a means to mitigate the adverse effects of soil contamination. The overall health of the soil plays a pivotal role in determining the extent to which contaminants may pose harm to human health.

Exposure assessments (NRC 1991; Swartjes 2015) are frameworks used to identify potential exposure pathways for humans for specific chemicals. An important part is evaluation of how contaminants move and persist and where they end up (e.g., in water, air, or soil) after they enter the environment. From this information, the most likely pathways for human exposure can be identified (Figure 6-4). There are multiple routes by which soil is involved in human exposure (NRC 1991; Swartjes 2015). For instance, individuals near contaminated areas can be exposed directly through dermal contact of the skin with soil or through ingestion of soil, which may occur from incidental ingestion, from consumption of soil particles remaining on foods, or through consumption of contaminants concentrated in foods and animal products. Apart from these direct health effects, contaminants may also be leached through the soil into groundwater and surface water, which can affect drinking water quality and accumulate in seafood. Finally, contaminants on the soil surface may be aerosolized, leading to exposure through inhalation. Whatever the route, estimates are made of contaminant concentrations in exposure media using monitoring data, environmental fate models, or both. These estimates are then used to gauge the potential contaminant intake by individuals, considering factors such as ingestion and inhalation rates (NRC 1991; Swartjes 2015).

FIGURE 6-4

An example of contaminant source, fate, transport, and exposure setting. NOTE: Leaking drums are the source of contamination in this figure. Chemicals are released into the air via volatilization, soil via leakage, and water via leaching. The chemicals (more...)

Bioavailability of Soil Contaminants

Bioavailability refers to the propensity of a contaminant to be taken up (e.g., absorbed by various biota, including plants, microorganisms, and human receptors. Bioavailability in soil is a multifaceted concept influenced by physical, chemical, and biological factors, as examined by the National Research Council (NRC 2003). These factors encompass a wide range of interactions that include soil properties (e.g., texture and pH), contaminant properties (e.g., chemical form and solubility), and biological factors, notably microbial activity. Bioavailability can vary over time due to evolving environmental conditions and processes. Considering bioavailability is paramount in the assessment of risks at contaminated sites and in the development of effective remediation strategies.

Barriers presented by soil and plants can diminish the transmission of contaminants to humans. The “soil–plant barrier” concept (Chaney and Ryan 1994; Basta et al. 2005) aids in discerning the relative significance of food chain versus soil ingestion exposure pathways, particularly for contaminants such as metals. The soil barrier, for instance, can reduce contaminant bioavailability through robust absorption mechanisms. Conversely, for the majority of plants, the plant barrier comes into play when contaminants exhibit phytotoxicity, potentially harming food crops before reaching levels detrimental to human health. Contaminants strongly adsorbed to soil, such as lead, encounter the soil barrier and do not pose a risk via the food chain; instead, ingestion of contaminated soil or dust becomes the primary risk driver. Conversely, highly mobile and weakly adsorbed contaminants such as cadmium bypass the soil–plant barrier, contaminating food crops and elevating the risk associated with the food chain pathway.

Soil properties and components, such as clay and organic matter content, exert substantial influence on contaminant solubility, mobility, and bioavailability. Soil pH, in particular, can modify contaminant bioavailability and mobility by regulating heavy metal dissolution/precipitation, by influencing the ionization of pH-dependent ion exchange sites on organic matter and metal oxide clay minerals, and by regulating strong adsorption of heavy metals and oxyanion metalloids (e.g., arsenate, arsenite) on soil clay minerals. Furthermore, pH-dependent functional groups on soil organic matter affect the formation of stable organometallic complexes. For organic contaminants such as pesticides, bioavailability varies, with nonpolar organic chemicals primarily adsorbed via partitioning to soil organic matter. Polar organic compounds, for example, perfluoroalkyl acids (a type of PFAS), exhibit anionic characteristics within the environmentally relevant pH range. The length of the hydrophobic chain controls the extent to which PFAS contaminants are adsorbed by soil organic matter.

The type of adsorption significantly shapes the environmental fate and long-term exposure of contaminants (Table 6-1). Significant advances have been made to characterize the type of metal contaminant adsorption to soil. Application of synchrotron spectroscopy including X-ray absorption near edge structure and extended X-Ray absorption fine structure for lead and arsenic have been reviewed (Sparks 2013). Limited studies have reported the relationship between metal speciation in soil and their bioavailability or bioaccessibility for lead or arsenic (Scheckel et al. 2009; Noerpel et al. 2020). Over the last two decades, in vitro gastrointestinal (IVG) soil extraction methods have been reported to be highly predictive of lead and arsenic bioavailability (Drexler and Brattin 2007; Basta and Juhasz 2014; Diamond et al. 2016; Whitacre et al. 2017). The IVG extraction methods have been shown to be more predictive of bioavailability than chemical speciation methods via synchrotron spectroscopy (Stevens et al. 2018). Weak adsorption processes, such as electrostatic and hydrophobic partitioning to organic matter, reduce bioavailability and mobility in the short term but not over the long term. In contrast, strong and irreversible sorption is desirable for persistent contaminants like heavy metals and PFAS, effectively reducing long-term human exposure.

TABLE 6-1

Bioavailability and Strength of Adsorption of Select Contaminants to Soil Adsorbent Phases.

In addition to adsorption reactions, metal contaminants can be sequestered as minerals or in soil minerals. Formation of double layer hydroxides (LDH) have been shown to sequester metals including cobalt, nickel, and zinc. Sequestration of these metals in LDH has been reviewed (Siebecker et al. 2018). This sequestration mechanism has not been reported for the metals of concern (lead, arsenic, and cadmium) in this chapter.

CONTAMINANT CASE STUDIES

In this section, the source, fate and transport, exposure pathway, and impacts of major classes of soil contaminants on soil ecosystems and human health are summarized. Although not exhaustive, these examples represent some of the most pressing soil contamination issues that pose a risk to human health. While contaminant sources and exposure routes are not exclusive to agriculture production, both the sources and solutions have, in many cases, a direct connection to agricultural management practices and soil health.

Heavy Metals and Metalloids

Many heavy metals and metalloids can be toxic to humans and the environment, depending on exposure routes, concentrations, and bioavailability. Lead and arsenic are ranked as the top two substances found at national priorities sites determined to pose the most significant potential threat to human health due to their known or suspected toxicity and potential for human exposure.¹ Lead is the most common heavy metal soil contaminant worldwide (Hettiarachchi et al. 2023). The environmental chemistry of lead (a cation) and arsenic (usually an oxyanion) are used here to demonstrate behavior and health impacts of these toxic trace elements. Cadmium is included as an example because it is a highly mobile toxic metal that concentrates in food crops.

Sources of Lead in the Soil Environment

Lead occurs naturally in soils, but most soil lead contamination is a consequence of anthropogenic activities. Mining, smelting, refining activities, and coal combustion have resulted in widespread contamination of soil (Rieuwerts et al. 1998; Tchounwou et al. 2012; ATSDR 2020). Leaded gasoline and paint are also major sources of soil contamination (ATSDR 2020). Ash deposits from solid waste incineration prior to the U.S. Solid Waste Disposal Act of 1965 have been found to be a source of lead contamination in soil as well (Bihari et al. 2023).

In agriculture, lead-containing inorganic pesticides such as lead arsenate (PbHAsO4) were extensively used because of its immediate effectiveness, low cost, and easy handling (Hettiarachchi et al. 2023). Lead arsenate was widely used in U.S. agriculture until the 1950s; its heavy use in orchards over many years caused the buildup of lead as well as arsenic in the soil environment (Schooley et al. 2008). Inorganic fertilizers, manure, compost, and biosolids are alternative sources of accumulation of lead, either in soil or plants (Alengebawy et al. 2021).

Sources of Arsenic in the Soil Environment

Similar to lead, arsenic is found naturally in soil, but most contamination of soil with arsenic is primarily related to anthropogenic sources. These sources include nonferrous metal mining and smelting, wood and coal combustion, wood preservation, pesticide application, and waste incineration (ATSDR 2007). Irrigation water containing high concentrations of arsenic can also be a significant source of soil contamination in India, Bangladesh, and east Asian countries (e.g., Vietnam, China) (Mitra et al. 2017). Arsenic contamination of rice, especially brown rice, from arsenic-tainted irrigation water has been reported in the U.S. Southeast (Meharg et al. 2009) and in California (Carrijo et al. 2022). Arsenic contamination in the southeastern soils is likely from the historical use of arsenical pesticides for cotton production before these pesticides were banned in the 1980s and 1990s; conversion of land from cotton production to rice production has resulted in elevated levels of arsenic in rice in the Mississippi Delta and Texas (Zavala and Duxbury 2008).

Sources of Cadmium in the Soil Environment

Cadmium is one of the most toxic heavy metals and negatively affects essential biological processes of humans, plants, and animals (Kabata-Pendias 2010). Major sources of cadmium pollution in soil include zinc mining, processing, and smelting and the use of phosphate fertilizer produced from rock phosphate ore with elevated cadmium (Kabata-Pendias 2010). Smelter emissions from zinc and cadmium production facilities have contaminated downwind environments (Chaney and Ryan 1994; Zhou et al. 2022). Land application of phosphate fertilizer can result in significant soil pollution with cadmium that impairs crop quality (Kabata-Pendias 2010). However, long-term application of phosphate fertilizer with low cadmium content does not increase soil cadmium or impair wheat grain quality (Basta et al. 1998). Cadmium can also reach soil through land-applied biosolids (ATSDR 2012).

Impact on Soil Ecosystems

Heavy metals change soil microorganism community structures and diversity (Konopka et al. 1999; Khan et al. 2010; Rodríguez Martín et al. 2014; Gutiérrez et al. 2016). Khan et al. (2010) determined that the high levels of cadmium and lead in soils caused decreases in microbial biomass carbon, inhibited acid phosphatase and urease enzymatic activity, and changed community structure based on denaturing gradient gel electrophoresis banding patterns. They also found that bacteria are more sensitive to heavy metal concentrations than actinomycetes and fungi. Konopka et al. (1999) found lead soil content to correlate with microbial biomass and determined that some bacteria populations contained lead-resistance genes. They noted that heavy metal–contaminated soils also have reduced plant biomass that could affect microbial community composition and activity. Others have noted that the diversity and function of soil invertebrates, such as earthworms and nematodes, can be altered by cadmium and lead (Žaltauskaitė and Sodienė 2010; Rodríguez Martín et al. 2014; Gutiérrez et al. 2016; Kavehei et al. 2018).

Fate and Transport: Human Exposure Pathways

Fate and transport of heavy metals is complex and varies greatly between metals. Three exposure pathways important to human health are: soil/dust ingestion and inhalation, ingestion of contaminated food, and ingestion of contaminated drinking water. The first two pathways, which relate to soil, are reviewed here.

Soil and dust ingestion

One of the major exposure pathways for lead, arsenic, and other heavy metals to humans is through the incidental ingestion of soil or dust. Ingestion is of special concern for children due to their increased hand-to-mouth activity and enhanced pharmacokinetics. Many heavy metals, including lead and arsenic, often have low water solubility in the environment. Low solubility reduces the transport of metals from soil to crops and from soil to source waters. The exception is cadmium, which escapes the soil–plant barrier and easily contaminates food crops (Basta et al. 2005). Many of the heavy metals in ingested dust or soil are dissolved in the acidic conditions (e.g., pH 1.5 to 2.5) of the upper gastrointestinal tract.

Soil properties and agricultural management practices can affect bioavailability and exposure to lead and arsenic. Lake et al. (2021) reported key soil properties reduced the bioavailability of arsenic by 17–96.5 percent and of lead by 1.3–38.9 percent associated with soil human ingestion. For both arsenic and lead, bioavailability decreased with increasing content of aluminum oxide and iron oxide. Soil amendments that add aluminum oxide and iron oxide to soil (e.g., via biosolids) will decrease arsenic and lead bioavailability.

Contaminant transport from soil to the food chain

In general, transport of many heavy metals from contaminated soil to plants to humans is small when compared to incidental ingestion (Chaney and Ryan 1994; Attanayake et al. 2015). However, small amounts of heavy metal absorption into crops that are diet staples (e.g., wheat, rice) can result in significant human exposure. For example, root crops such as carrots, radishes, and beets can accumulate soil lead, and surface contamination of crops with soil can affect human exposure (Attanayake et al. 2014, 2021). Human exposure from ingestion of soil adhered to vegetables has been found to be greater than amounts in those vegetables. Best management practices to reduce heavy metal exposure include washing crops thoroughly before consumption to get rid of adhering soil particles.

Diet is the largest source of exposure to arsenic, largely from grains, produce, seafood, and drinking water. Rice, in particular, is an important source of dietary arsenic because it is cultivated under flooded conditions leading to an anoxic soil environment that liberates arsenic from its minerally bound form and allows it to be taken up and concentrated in the grain. Notably, a recent study found that arsenic availability, uptake, and allocation in rice increases in warmer temperatures; therefore, global warming may increase the risk of arsenic exposure through consumption of rice in production systems that were previously considered low risk (Farhat et al. 2021).

For non-smokers, diet is the largest exposure route for cadmium (ATSDR 2012). However, the concentration of cadmium in the soil is not predictive of the concentration of cadmium in food grown in the soil; soil properties, the type of crop and its genetics, crop rotation, and fertilizer and irrigation management practices all affect the ultimate concentration of cadmium in food (Schaefer et al. 2020).

Impacts on Human Health

Lead exposure primarily occurs through incidental ingestion of contaminated soil and dust, with children being particularly vulnerable. Unlike respiratory and dermal pathways, incidental ingestion involves unintentional swallowing due to hand-to-mouth contact. Studies have consistently highlighted the significance of this pathway, with exposure typically occurring through hand-to-mouth transfer of contaminated soil or dust (Mielke and Reagan 1998; Glorennec et al. 2010; Oulhote et al. 2011; Zahran et al. 2013; Henry et al. 2015). Once lead is absorbed into the body and enters the bloodstream, it can rapidly distribute to the kidneys and liver or be slowly absorbed by soft tissues. The majority of the body’s lead burden is stored in teeth and bones. Lead is highly toxic and can affect nearly every organ, especially the nervous system. At elevated exposure levels, lead can lead to comas, convulsions, and even fatalities. There are no safe blood lead levels, and any concentration of lead can be associated with decreased IQ levels in children and behavioral challenges. These effects are considered permanent. Additionally, lead exposure can cause symptoms such as dizziness, irritability, fatigue, impaired cognition, anemia, hypertension, kidney damage, and reproductive organ toxicity (Wani et al. 2016; ATSDR 2020). Lead exposure during pregnancy has been linked to maternal health risks, including hypertension and pre-eclampsia. One study found that women living in areas with higher soil lead levels were more likely to suffer an eclampsia event (Zahran et al. 2014). Even low levels of exposure during pregnancy have been found to lead to poor birth outcomes including low birthweight and preterm birth (Bellinger 2005).

Inorganic forms of arsenic like As(III) (i.e., arsenite) and As(V) (i.e., arsenate) are associated with more severe health effects from exposure as compared exposure to organic arsenic. Ingestion of arsenic can cause gastrointestinal symptoms, decreased production of white and red blood cells leading to bruising, impaired nerve function leading to burning sensations in the hands and feet, and skin changes (Mandal and Suzuki 2002; ATSDR 2007). The U.S. Environmental Protection Agency (EPA) and the International Agency for Research on Cancer have also listed arsenic as a human carcinogen because it has been linked to cancer of the liver, skin, bladder, and lung. Fetuses, infants, and children are particularly vulnerable to the potential harmful effects from arsenic exposure, and exposure during times of active brain development has been linked to learning disabilities, behavioral problems, and lowered IQ.

As described above, the impact on human health can vary based on the timing duration, route of exposure, and host variability. Notably, recent research has shown that gut microbiota can increase the bioavailability of soil-bound arsenic thereby enhancing its toxicity (Yin et al. 2017; McDermott et al. 2020; Griggs et al. 2022). While research in humans is limited, studies using in vitro models such as the Simulator of the Human Intestinal Ecosystem have shown that human gut microbiota may increase the bioavailability of arsenic in contaminated soil samples (Laird et al. 2007; Van de Wiele et al. 2010). Despite this individual variability, the U.S. Food and Drug Administration has issued guidance to industry on acceptable levels of arsenic in foods often consumed by infants and children. It recommends that inorganic arsenic levels do not exceed 100 parts per billion (ppb) in infant rice cereal (FDA 2020) and 10 ppb of inorganic arsenic in single-strength (ready-to-drink) apple juices (FDA 2023).

Recent research has also described the impact of heavy metals and metalloids on the gut microbiome, which is important for human health. Studies using animal model systems have shown significant alterations of gut microbiomes to arsenic exposure (K. Lu et al. 2014; Gokulan et al. 2018). One study in mice showed that arsenic exposure not only caused shifts in the composition of the gut microbiota but also led to impairment of the immune response and higher inflammation. Further, these changes were dose-dependent and differed according to the age of the animals (Gokulan et al. 2018). The function of the gut microbiome also is impacted by arsenic exposure, which in one study inhibited the fermentation rate of rumen bacteria by almost one third (Forsberg 1978).

Information about the effect of arsenic on the human microbiome is scarce, but vast shifts in the composition of gut microbial communities have been observed in populations exposed to arsenic (McDermott et al. 2020). For example, in Bangladesh, there was a significantly higher abundance of Proteobacteria as well as multidrug and arsenic resistance microbial genes in children exposed to arsenic than their unexposed counterparts (Dong et al. 2017). In the United States, urinary arsenic was correlated with several taxonomic groups in naturally exposed infants, although none of them were in the Proteobacteria phylum (Hoen et al. 2018). Significant compositional shifts were also observed in a mining community in China with prolonged exposure to heavy metal contamination when compared to a community with little exposure to these pollutants. Paradoxically, metal exposure was also correlated with microbial richness in the human gut (Shao and Zhu 2020).

Evidence supports adverse effects on the kidneys and bone density from chronic dietary cadmium exposure, although in some studies, the populations also lived in areas polluted with cadmium. Studies of the effects of dietary cadmium exposure on human health have found mixed results on cardiovascular disease. Consumption of cadmium at high concentrations is known to cause gastrointestinal distress (ATSDR 2012).

Mitigation and Remediation

Remedial approaches for lead are based on removal or reducing the bioavailability of lead. Bioavailability-based remediation is a cost-effective in situ approach that does not remove the contaminant but does reduce the contaminant’s bioavailability. The use of soil amendments to remediate lead-contaminated soils focuses mainly on changing existing soil lead chemistry in situ. It is done by inducing the formation of sparingly soluble lead solids or enhancing chemisorption. Many soil amendments have been used successfully for bioavailability-based remediation (Henry et al. 2015; Wang et al. 2021; Hettiarachchi et al. 2023). Commonly used soil amendments reported to reduce lead bioavailability are summarized in Table 6-2. Other amendments are in development including formation of plumbojarosite (Sowers et al. 2023) and the use of modified biochars (Yang et al. 2019).

TABLE 6-2

Commonly Used Soil Amendments for Bioavailability-Based Remediation of Lead-Contaminated Soil.

Several of the soil amendments are applied to agricultural fields through additions of phosphorus fertilizer, manures, compost, biochar, and liming materials. These amendments decrease lead bioavailability and exposure to human and ecosystem receptors. Many amendments are green and sustainable remediation (GSR) strategies, which are holistic approaches that maximize environmental, social, and economic benefits (Wang et al. 2021).

Phytoextraction is using hyperaccumulator plants that absorb metals (e.g., more than 10,000 mg/kg) from soil followed by harvesting of the hyperaccumulator plants. Phytoremediation of cadmium, zinc, selenium, and arsenic from contamination soil has been reviewed (Chaney et al. 2014). Hyperaccumulator plants for lead do not exist. Application of chelating ethylenediaminetetraacetic acid solution to lead-contaminated soil planted with Brassica juncea (e.g., Indian mustard) was able to accumulate more than 1 percent of the lead in plant tissue (Blaylock et al. 1997). However, this practice contaminated groundwater (Nowack et al. 2006).

Remediation strategies vary depending on the form of arsenic in the environment. Often, the dominant form of arsenic contamination in aerobic soil is arsenate. The majority of in situ remediation methods are designed to adsorb or precipitate arsenate and reduce its bioavailability and mobility (Gong et al. 2018). In particular, iron oxide clay minerals strongly adsorb arsenate and reduce arsenate bioavailability (Violante et al. 2010). More recently, GSR soil amendments for remediation or arsenic-contaminated soils include biochar (Wang et al. 2021) and biochar modified with iron oxide (Yang et al. 2021). These treatments improve soil health while remediating arsenic.

Mitigating cadmium in the food supply is complicated by the properties of the soil the food is grown in, the type of amendments that may be applied to the crop, and the genetics of the crop. Cadmium can also be introduced to the food supply from utensils, plastics, and cookware that contain cadmium. When it comes to steps to reduce cadmium uptake in the field, better understanding of geogenic sources and specific plant uptake would help mitigate cadmium in food, as would minimizing the application of phosphate fertilizers and irrigation water that contains cadmium (Schaefer et al. 2020). Producers would benefit from research that provides evidence for management strategies tailored to specific soils and specific crops that prevents plant uptake of cadmium (Schaefer et al. 2020).

Microplastics

Microplastics (MPs) are defined as small plastic particles measuring less than 5 mm in size (Masura et al. 2015). MP pollution includes primary MPs (e.g., microfibers, textiles) and secondary MPs, which are defined as degradation products of larger plastic products (Petersen and Hubbart 2021). Approximately 90 percent of synthetic polymers produced are polyethylene (PE), polypropylene (PP), polyvinyl chloride (PVC), polystyrene (PS), or polyethylene terephthalate (PET) (Andrady and Neal 2009). MPs exist in a variety of physical forms that include fibers, foam, film, and particles (Rillig et al. 2019a). Found in diverse ecosystems worldwide, MPs have become a pervasive and global environmental issue (Bolan et al. 2020; Büks and Kaupenjohann 2020) and a major reason the planetary boundary for the category of contaminants (i.e., novel entities) is being exceeded (Persson et al. 2022).

Sources

Microplastics reach soil through direct and indirect pathways, which vary based on location, land use, and management practices (Figure 6-5; Yadav et al. 2022). In agriculture, the primary sources are compost (Bläsing and Amelung 2018), biosolids (Bläsing and Amelung 2018; Crossman et al. 2020), irrigation with untreated wastewater (Bläsing and Amelung 2018; He et al. 2018), and polymer coatings of slow-releasefertilizers (Lian et al. 2021). Worldwide, plastic mulch use in agricultural production is a growing source as well (He et al. 2018).

FIGURE 6-5

The source, movement, and fate of microplastics in the environment. SOURCE: Used with permission of Springer Nature BV, from “Unravelling the Emerging Threats of Microplastics to Agroecosystems”, Yadav et al., Reviewers in Environmental (more...)

Once in soil, MPs tend to accumulate near the soil surface due to their low density and small particle size. MPs can also become incorporated into soil aggregates or immobilized on to soil particles as a function of the particular soil’s texture, organic matter content, and moisture levels (Rillig and Lehmann 2020). MPs can be transported within and out of soil through processes such as surface runoff, erosion events, and leaching to groundwater and via biological activities such as earthworm burrowing and root growth.

Fate and Transport

The persistence and degradation of plastics and MPs in soil environments have been reviewed by Restrepo-Flórez et al. (2014), Krueger et al. (2015), and Bläsing and Amelung (2018). Plastics, synthesized to be durable and withstand deterioration, accumulate in soil (Bläsing and Amelung 2018) despite being susceptible to degradation by abiotic and biotic processes (Figure 6-5). Degradation rates are dependent on the type of MPs, location in soil, management practices, climate, and other factors. Degradation pathways via ultraviolet (UV) radiation are known, particularly in aquatic ecosystems, but in soil only materials near the surface are susceptible and degradation rates are generally slow. Over time, UV exposure and weathering processes cause larger MPs to fragment into smaller particles, which then move through soil and become more susceptible to both uptake and degradation (Hooge et al. 2023).

With respect to biodegradation, a microbial process, most plastics degrade very slowly (if at all) in soil, with persistence estimated to range from several years to several thousand years (Chamas et al. 2020). The usually low surface-to-volume ratio of most forms of microplastics physically limits their bioavailability to microorganisms, who need to maintain close proximity to surfaces for reactions (Hooge et al. 2023). Field studies reviewed by Bläsing and Amelung (2018) indicate minimal soil degradation of MPs with far less than 1-percent loss by weight of PE over 2.5 year (Albertsson 1980) and PP after 1 year (Arkatkar et al. 2009) and no detectable PVC degradation after 10–35 years in soil (Otake et al. 1995; Santana et al. 2012; Ali et al. 2014). Thus, although large and visible pieces of plastics may fragment over time, most data confirm that MPs are persistent in soil (Krueger et al. 2015; Bläsing and Amelung 2018). Unfortunately, rates of inputs of MPs greatly exceed their rates of removal in soil (Petersen and Hubbert 2021).

Impact on Soil Ecosystems

Physical–chemical

MPs are redistributed throughout the soil by tillage, by physical mixing from wet-dry or freeze-thaw cycles, or by soil organisms (Hooge et al. 2023). MPs can affect soil structure and porosity, leading to changes in water infiltration, retention, and drainage (Figure 6-6; de Souza et al. 2019; Lehmann et al. 2019) and, in turn, influence nutrient availability, root growth, and plant productivity. Due to their larger size (around 2 mm), fiber, foam, and film MPs can potentially increase soil aeration and microporosity (Sun et al. 2022). MP microfibers can decrease the water stable macroaggregates via both physical effects and biological effects on soil biota that contribute to soil aggregation (Boots et al. 2019). MPs can also modify soil chemical properties by adsorbing and releasing contaminants, such as heavy metals and organic pollutants, thereby influencing their bioavailability and potential toxicity to soil organisms and plants (Sajjad et al. 2022).

FIGURE 6-6

A schematic diagram of microplastic interaction with plant–soil systems. SOURCE: Used with permission of Springer Nature BV, from “Unravelling the Emerging Threats of Microplastics to Agroecosystems”, Yadav et al., Reviewers in (more...)

Impacts on soil communities

MPs have been documented to change the diversity, abundance, and functions of soil microbial communities (Rillig et al. 2021b); the extent of impact is often proportional to concentration (Sun et al. 2022). Overall effects on soil biota can be negative, positive, or neutral and occur via both direct (e.g., absorption of MP nanoparticles) and indirect (e.g., altering soil structure) pathways. Though not well understood, the impacts are many (Rillig et al. 2021b), such as shifting the composition of microbial communities and altering microbial processes, including carbon storage, organic matter decomposition, and greenhouse gas emissions (Rillig et al. 2021a,b). MPs are made up primarily of recalcitrant carbon and can slow down overall organic matter decomposition rates (Rillig et al. 2021a). MP fibers in soil resulted in an increase in CO₂ fluxes and a decrease in nitrous oxide emissions, especially after urea fertilizer application, which is attributed to increased soil aggregation and air permeability (Rillig et al. 2021a).

The plastisphere—defined as the soil surfaces influenced by plastic—has a different microbial community than bulk soil. For example, more pathogens and antibiotic resistance genes are present as compared to bulk soil (Zhu et al. 2022; Rillig et al. 2024). Furthermore, plastics have been found to change soil bacteria community structure and interfere with microbial lipid metabolisms and the biosynthesis of secondary metabolites (Wu et al. 2022). MPs can detrimentally affect the growth, reproduction, lifespan, and overall survival of soil fauna through various mechanisms, including ingestion, bioaccumulation, oxidative stress, reproductive and neurotoxic effects, metabolic disruptions, and gut microbiota imbalances (Wang et al. 2022). Some soil fauna play a role in the formation and degradation of MPs, modify their movement in soil, and can potentially transfer accumulated MPs up the food chain (Helmberger et al. 2020; Wang et al. 2022).

The indirect effects of MPs on plant growth include alteration of soil structure, bulk density, and water-holding capacity, which in turn influences plant development (Helmberger et al. 2020). These effects are varied and can be both positive and negative, heavily dependent on the characteristics of the MPs (like their chemical additives) and the specific types of soil and plants in the environment (Rillig et al. 2021b). MPs can impair plant seed germination, reduce root elongation, and negatively affect plant biomass and reproductive capacity (Boots et al. 2019; de Souza et al. 2019). MPs may also interfere with plant nutrient uptake and induce oxidative stress, leading to physiological disorders and decreased overall plant performance (Rillig et al. 2019b). These impacts could have a cascading effect on other biophysical processes in the soil and may pose a threat to soil biodiversity and ecosystem functioning (Boots et al. 2019).

Human Exposure Pathways

MPs in soils can pose potential risks to human health through several exposure pathways such as direct soil ingestion, particularly by children during outdoor activities as well as inhalation of MP particles suspended in the air due to wind erosion or dust generation from contaminated soils. Consumption of contaminated foods is also possible; most research to date has focused on MPs in food that is not directly related to soil, such as seafood, fish, salt, honey, and water (Karbalaei et al. 2018). One study that investigated the quantity of MPs in fruits and vegetables found that the median distribution of MPs <10 micron in size ranged by 52,050 to 233,000 particles/g and that apples were the most contaminated samples (Conti et al. 2020). The estimated daily intakes associated with that study are smaller than the estimated daily intakes from plastic bottled mineral water (Zuccarello et al. 2019), therefore the committee concluded that consumption from soil-related agricultural produce may be less of a concern.

Impacts on Human Health

The human health effects of MPs can be categorized into chemical, physical, and biological effects, as illustrated in Figure 6-7. Chemical effects refer to the impact of additives and dyes that may be toxic, teratogenic, or carcinogenic. For example, phthalates are commonly used as plasticizers to provide flexibility to plastics. Because they are additives, they are not chemically bound (covalently bonded) and are more likely to be released and transferred to the environment (Blackburn and Green 2022). Similarly, polybrominated diphenyl ethers are used as flame retardants in many commercial products and, because they are not chemically bound, can leach during production and recycling, bioaccumulate, and cause endocrine disruption and impaired neurological development in animal models. Secondary toxins include persistent organic pollutants that can lead to immunotoxicity and secondary toxicity through interaction with other chemical pollutants. For example, plastic mulching sheets used in agriculture can both absorb pesticides and increase risks for human exposure (Huang et al. 2020, Wang et al. 2020).

FIGURE 6-7

A flow diagram illustrating the potential human health effects of microplastics. NOTE: The dotted lines represent current speculative research. SOURCE: Used with permission of Springer Nature BV, from “The Potential Effects of Microplastics on (more...)

Physical effects include inhalation and ingestion of MPs and subsequent immunotoxicity. Small MPs are common in the atmosphere and can be readily inhaled; however, the health effects remain poorly understood. Eyles et al. (2001) found that, once inhaled or ingested, MPs could also translocate to other tissues; fluorescent polystyrene microspheres delivered intranasally to mice were found in the spleen 10 days later. MPs have also been shown to impair gastrointestinal function and alteration of hepatic lipid metabolism in animal models (Y. Jin et al. 2019; L. Lu et al. 2019); this may also be the case for humans.

Finally, there are biological effects that primarily refer to the impact on the microbiome. Evidence thus far, mostly stemming from animal models, has shown that nano and microplastics alter microbial community structures in the gut (both alpha and beta diversity), which may in turn underlie immune impairment, as previously reviewed (Fackelmann and Sommer 2019; L. Lu et al. 2019; Santos et al. 2022).

As an additional burden, microplastics from soil may carry pathogens and other pollutants that pose a threat to human health potentially through alterations in gut microbial communities (Fackelmann and Sommer 2019). A mice study found that when co-ingested with polyethylene particles, the bioavailability of arsenate increased, and this seemed to be mediated by the metabolite output of the gut microbiota (Chen et al. 2023). Also, impacts of MPs on soil communities may translate into effects on the food system and have negative consequences for human health (Daghighi et al. 2023).

Few studies have comprehensively assessed the impact of microplastics on health concurrently with their presence in the environment. A recent study compared the soil, air, nasal, and gastrointestinal microbiomes of two populations, one with high exposure to microplastic pollution and the other with low exposure. They found overlap in the microplastic species found in gut microbiome samples and soil samples in the high exposure area, along with inverse correlations between microplastic exposure and the abundance of taxonomic groups that may be beneficial for human health (Zhang et al. 2022).

Mitigation and Remediation

The reduction of plastic use in agriculture is an essential step toward more sustainable and environmentally friendly farm management. To address these issues, the “3 R” waste hierarchy concept of reducing, reusing, and recycling plastics, before considering disposal, is crucial (Hofmann et al. 2023). Similarly, the United Nations’ Food and Agriculture Organization (FAO) has promoted the 6R model (Refuse, Redesign, Reduce, Reuse, Recycle, and Recover) for reduction of use of plastics in agriculture (FAO 2021). Both approaches advocate for plastic applications that have circular end-of-life treatment options and stress the importance of innovative material design to ensure plastics can be completely collected, recycled, and reused. Use of non-biodegradable polymer-coated fertilizers and mulching films, in particular, should be minimized and, if used, be collected after use. Additionally, in situations where plastics cannot be collected after use, or where recovered plastic is too degraded or soiled to be reused or recycled, polymers that are more biodegradable or less toxic should replace conventional persistent polymers (Galati and Scalenghe 2021; Hofmann et al. 2023).

Biodegradable polymers offer potential solutions for reducing environmental impacts of plastics for short-term uses (e.g., cutlery). Biodegradable plastics can break down through composting or exposure to UV radiation. However, uncertainties concerning biodegradable polymers involve their complex waste management, including requirement for specific collection and composting facilities, and the low volumes produced that may not justify the waste management efforts (Prata et al. 2019). Additionally, some degradable plastics produce non-degradable by products (Prata et al. 2019).

The possibility of significant and consistent biodegradation of conventional plastics in the environment remains uncertain (Krueger et al. 2015). Some strains of bacteria and fungi have been demonstrated in the laboratory to be capable of degrading many polymers via enzymatic hydrolysis or oxidation (Sivan 2011; Krueger et al. 2015). Some conventional plastics can biodegrade under lab conditions when exposed to specific plastic-degrading organisms, such as Zalerion maritimum (Paço et al. 2017); how successful these organisms would be in soil is not known.

PFAS

Per- and polyfluoroalkyl substances are a large group of synthetic, organofluorine chemicals that are of increasing concern as environmental contaminants. Their characteristic carbon–fluorine bond is one of the strongest single bonds in chemistry, so PFAS molecules are extremely hard to break down. Thus, they are highly persistent in the environment and are often referred to as “forever chemicals.” There are thousands of PFAS compounds² of concern with respect to the health of people and the environment.

PFAS give desirable properties to industrial and consumer products because they resist water, oil, and heat. They are used to make waterproof and stain-resistant garments, nonstick pans, and oil-resistant containers. Teflon is a well-known example. They are also used to make firefighting foams, fabrics that resist fire and stains, and a great many other products and industrial processes.

The different PFAS species vary in their length of the carbon chain, their branching structure, the number of fluorine molecules and their positions, attached functional groups, and so on (Buck et al. 2011). The large and diverse PFAS family of molecules can be divided into the polymer and non-polymer classes, with the latter most commonly found in biological and environmental samples. This non-polymer class includes per- and polyfluoroalkyl substances; the carbon chains of the perfluoroalkyl substances are fully fluorinated, while the carbon chains of the polyfluoroalkyl substances are not fully saturated with fluorine atoms. The perfluoroalkyl acids (PFAAs) are one major group of the perfluorinated subclass of the non-polymer class of PFAS. Because of their saturation, they are the most stable and thus most persistent in the environment. Their non-saturated precursors can be transformed into PFAAs through biological or chemical cleavage of their non-fluorinated moieties.

Sources

PFAS molecules are created through industrial processes. They have been manufactured since the 1940s and widely used since the 1950s. They are spread from both point sources (where they are made or used) and diffuse sources (such as water, soil, and air). Point sources include industrial sites where PFAS is made and used, such as fluorochemical production plants (Gebbink and van Leeuwen 2020).

Several studies have found soil to be a major environmental reservoir of PFAS (Strynar et al. 2012; Brusseau et al. 2020). Soils can receive PFAS from applied biosolids, contaminated irrigation water (Pepper et al. 2021), aqueous film-forming foam used extensively at military bases and airports (Yan et al. 2024), and rainfall (Pfotenhauer et al. 2022). PFAS in biosolids can come from many sources, including landfill leachate. Landfill leachate is a rich source of PFAS because of the discarded domestic products (e.g., cookware, clothing, carpets, and furniture) and industrial products deposited in these sites (Lang et al. 2017; Capozzi et al. 2023). Landfill leachate is typically treated at the municipal wastewater treatment plants where other wastes, including excreta and industrial wastes, are processed (Masoner et al. 2020; Helmer et al. 2022).

Because the biological and other treatment processes implemented through conventional wastewater treatment do not destroy PFAS, biosolids (the stabilized solids that result from municipal wastewater treatment, also known as sewer sludge) may be contaminated. Longer-chain PFAS molecules (e.g., those with an 8-carbon backbone or C8) are typically enriched in biosolids relative to liquid effluent (Helmer et al. 2022).

Fate and Transport

PFAS are highly mobile in the environment (Brunn et al. 2023). As noted above, their forms can change in the environment as non-fluorinated moieties are cleaved, yielding stable (saturated) PFAAs (Evich et al. 2022). Greater amounts of some PFAS molecules are thus found in the effluent from wastewater treatment plants than in the influent (Coggan et al. 2019; Thompson et al. 2023). Short-chain PFAS are more mobile in soils than longer-chain forms (Brusseau et al. 2020). General pathways are indicated in Figure 6-8.

FIGURE 6-8

A diagram illustrating the sources of PFAS and pathways for their release into the water cycle and the environment. SOURCE: Diagram courtesy of Michigan Department of Environment, Great Lakes, and Energy.

Biosolids have long been used as a soil amendment, adding carbon and nutrients to agricultural and other lands (Lu et al. 2012). While it is desirable to regard organic wastes as resources, concern is rising about PFAS contamination of biosolids leading to contamination of farmland (Lowman et al. 2013; Mason-Renton and Luginaah 2018). Plants can take up PFAS, with short-chain forms showing greater mobility in plant tissues than longer-chain forms (Costello and Lee 2020). Maine recently banned the land application of biosolids after some farms were found to have high levels of contamination (Perkins 2022).

Impact on Soil Ecosystems

Studies of the effects of PFAS on soil microbiomes have shown that PFAS changes community composition and metabolite production (Senevirathna et al. 2022; Xu et al. 2022; Wu et al. 2023). Several studies have reported that PFAS reduce microbial abundance, including in laboratory experiments (Xu et al. 2022) and field studies comparing contaminated and noncontaminated sites (Senevirathna et al. 2022). Xu et al. (2022) found that soil treated with the PFAS compounds perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) had bacterial reduced gene abundance as compared to the control treatment and that PFAS exposure altered soil microbial composition. The researchers speculated that the increase in Proteobacteria was due to Proteobacteria’s greater tolerance for PFAS (Xu et al. 2022). Increases in Proteobacteria have also been observed in other studies (Wu et al. 2023). Modeling also predicted that exposure to PFAS inhibited numerous microbial metabolism processes in soil. Wu et al. (2023) found that soil microorganisms downstream of a Teflon production plant had reduced lipid biosynthesis and possibly decreased metabolic activity.

Findings with regard to effects of PFAS on soil pH and microbial diversity have varied across studies. Xu et al. (2022) did not find a change in soil pH between soil treated with PFOA and PFOS and a control sample, while Xu et al. (2023) found a significant increase in soil pH in another laboratory experiment. The latter results were likely because PFAS increased litter decomposition rather than PFAS having a direct effect on soil pH. The latter researchers also found that soil aggregate stability was not affected by PFAS but water-stable aggregates decreased.

The direction of change in microbial diversity and richness of soil microorganisms exposed to PFAS is unclear, possibly due to differences in PFAS concentration and species across studies. Xu et al. (2022) found that microbial diversity and richness of soil microorganisms increased in the treatment cases as opposed to the control in a laboratory study, whereas Senevirathna et al. (2022) observed a decline in soil bacterial community population and diversity in soils collected from contaminated sites versus uncontaminated environmental samples. PFOS (a C8 molecule) had a greater effect than PFOA (a molecule with a 7-carbon backbone), suggesting the type of PFAS may influence impacts on soil microbial communities. Cai et al. (2019) also found that the type of PFAS influenced its effect on microbial diversity and richness, with longer chains having greater toxicity than shorter chains.

Additionally, PFAS may negatively affect plant growth. Xu et al. (2022) found that PFOA and PFOS inhibited the abundance of Azospirillum, a plant growth-promoting rhizobacteria. PFOS also enriched Hydrogenophaga, which are resistant to organic pollutants, whileMethyloversatilis was more abundant under PFOA exposure. The abundance of Methyloversatilis suggests that it may be able to degrade PFOA. Wu et al. (2023) also determined that PFAS contamination affected the fungal community in soils. The authors noted that mycotoxin concentrations were higher downstream of the Teflon plant.

Human Exposure Pathways

People are exposed to PFAS through many pathways, including via occupational exposures, consumer and household products, environmental contamination, and ingestion. Occupational exposures are myriad and may include occur where PFAS-containing products are manufactured, in the construction industry, among food workers that handle PFAS-containing food packaging, and among firefighters. PFAS are used in thousands of consumer products ranging from personal care products like sunscreen, makeup, and dental floss to textiles, artificial turf, and paint. A recent review found that exposure from contaminated household dust could account for up to 25 percent of blood concentrations but that there were many methodologic flaws in the existing studies and that more research on household exposure pathways is needed (DeLuca et al. 2022).

Environmental contamination of soils and groundwater is also thought to be an important source of PFAS exposure. Most research in nonoccupational settings has focused on ingestion of PFAS, which can occur through drinking contaminated water or eating contaminated seafood, vegetables, game, or dairy products (Domingo and Nadal 2017; Death et al. 2021). In April 2024, EPA established a regulation for maximum contaminant levels for six PFAS compounds for public water systems in the United States (EPA 2024a). It estimated that 6–10 percent of the country’s 66,000 public drinking water systems would have to take action to reduce PFAS to meet these levels (EPA 2024b). PFAS may also be ingested through food contaminated by PFAS-containing materials including food packaging, microwave popcorn bags, and cookware. PFAS can be transferred intergenerationally through the mother’s body to a developing fetus and through breastfeeding (Manzano-Salgado et al. 2015; Gao et al. 2019; Zheng et al. 2021). Exposure pathways may also include inhalation of aerosolized or volatile PFAS, which have been detected indoors and near factory emissions. The impact of inhaled and transdermal exposures, for example through bathing in contaminated water, have been less well studied (Sunderland et al. 2019).

Impacts on Human Health

PFAS are known to affect human health in multiple ways (Sunderland et al. 2019; Brase et al. 2021; Chambers et al. 2021). A recent review by the National Academies summarized the health effects of PFAS and found that cancers, endocrine effects, dysregulation of immune function, and impacts on fertility were the health effects most frequently mentioned by speakers at that committee’s information-gathering meetings (NASEM 2022a). After a systematic synthesis of the evidence, that report concluded that there was sufficient evidence to support an association with decreased antibody response (in adults and children), dyslipidemia (in adults and children), decreased infant and fetal growth, and increased risk of kidney cancer (in adults). The report also found suggestive evidence for the following diseases and health outcomes: increased risk of breast cancer (in adults), liver enzyme alterations (in adults and children), increased risk of pregnancy-induced hypertension (gestational hypertension and pre-eclampsia), increased risk of testicular cancer (in adults), increased risk of thyroid disease and dysfunction (in adults), and increased risk of ulcerative colitis (in adults).

Mitigation and Remediation

As PFAS have been found in drinking water, foods and food packaging, and the bodies of people and animals around the world, awareness and concern become increasingly palpable. This increased awareness has led to pressure for companies to stop making these compounds. After half a century of production, companies are beginning to phase out production of certain PFAS compounds. 3M began phasing out production of PFOS in 2000 (EPA 2000) and recently paid $10.3 billion to settle a multidistrict lawsuit (Kluger 2023). The company stated in 2022 that it would work to cease manufacturing PFAS by the end of 2025 (3M 2022). Reduced production has led to some decrease in the human body burden in the United States and Australian populations (Gomis et al. 2017). Although some PFAS forms are being phased out, there is evidence that companies are making replacement compounds with likely hazardous properties (Brase et al. 2021).

EPA recently published a roadmap related to PFAS contamination (EPA n.d.). As part of the roadmap, EPA is conducting a risk assessment for two specific PFAS compounds, PFOA and PFOS, in biosolids. The risk assessment is scheduled to be completed by December 2024.³ Given the large number of PFAS molecules, their persistence and potential to accumulate in the environment, their individual and collective hazards, and the difficulties of regulating them individually, some authors argue that they should be regulated as a class (e.g., Kwiatkowski et al. 2020). Others suggest that, because of the differences in toxicity among PFAS species, each PFAS subgroup and compounds within should be evaluated (Singh and Papanastasiou 2021). Cousins et al. (2020) argued that a rational approach to phasing out PFAS production would eliminate most applications but would recognize certain “essential uses.” Eliminating most uses but allowing exceptions for the most essential, such as certain medical applications, would likely be the most feasible approach.

Because of the extreme strength of the carbon–fluorine bonds that are the defining feature of PFAS, it is notably difficult to fully mineralize PFAS (that is, to break PFAS down to their elemental components rather than to smaller species of PFAS) (Shahsavari et al. 2021). Sorption of PFAS can block the movement of contaminants such as PFAS in the ecosystem. Porous, high-carbon materials such as activated carbon and biochar can be used to remove PFAS from drinking water (Box 6-1; Xiao et al. 2017). Granulated activated carbon is the most widely used sorbent for purification of air and water.Fecal biochar (e.g., made from biosolids) has been shown to be an effective sorbent for PFAS (Krahn et al. 2023).

BOX 6-1

Biochar as Decontamination Possibility.

That said, it is possible to degrade PFAS in soil and in amendments that might reach soil. A few alternatives are emerging, each with potential to address the issue in certain niches. Mechanochemical treatment by ball milling can be used to eliminate PFAS (Turner et al. 2021); this might be useful for extremely contaminated soils, though the committee finds it is difficult to envisage this being conducted at large scale in an economically viable way. Although bioremediation is challenging and generally not considered as a practical approach to PFAS remediation, there is hope that microbes can contribute to breaking down PFAS (LaFond et al. 2023). Another new biological approach involves use of enzyme-catalyzed oxidative humification reactions carried out by fungal extracellular enzymes (e.g., peroxidases and laccases). These processes can potentially lead to break down of certain PFAS molecules (Grgas et al. 2023; Kumar et al. 2023). Thermochemical transformation is an approach that has shown promise under certain conditions (Weber et al. 2023). Incineration can destroy PFAS but can lead to the emission of fluorinated byproducts (Stoiber et al. 2020).

There is evidence that pyrolysis at high temperatures (over 600°C) can be effective for removing PFAS from biosolids. Thoma et al. (2022) showed that 21 PFAS compounds detected in biosolid samples (ranging from 2 μg/kg–85 μg/kg) were undetectable in the biochar resulting from pyrolysis at 650°C. Kundu et al. (2021) demonstrated greater than 90-percent removal of PFOA and PFOS from sewage sludge after pyrolysis at temperatures between 500 and 600°C. McNamara et al. (2023b) showed more than 99-percent removal of target PFAS and PFAS precursors at pyrolysis temperatures up to 800°C. However, PFAS may be present in byproducts of pyrolysis (McNamara et al. 2023a,b). Further research and development is needed to assess the potential of pyrolysis as a general solution to the problem of PFAS in biosolids (Wallace et al. 2023). Other high-temperature treatments, plasma, and combustion under specific conditions can also be effective at breaking down PFAS (Singh et al. 2021; McNamara et al. 2023a,b; Weber et al. 2023). High-energy electron beam technology is also being tested to break down PFAS in soils and groundwater (Lassalle et al. 2021).

Integrated strategies for dealing with sources and flows of contamination are needed. These are likely to bring together elements of the strategies mentioned above, as well as selective mobilization of PFAS (Bolan et al. 2021). Figure 6-9 presents an overview of elements of remediation strategies that can be integrated.

FIGURE 6-9

Elements of a remediation strategy for PFAS. SOURCE: Reprinted from Journal of Hazardous Materials, 401, Bolan et al., “Remediation of Poly- and Perfluoroalkyl Substances (PFAS) Contaminated Soils–To Mobilize or to Immobilize or to Degrade?” (more...)

CONCLUSIONS

The properties of soil facilitate the capture and remediation of many contaminants. In some cases, the soil microbiome can mitigate the risk contaminants pose, such as through the attenuation of heavy metals via sequestration, ion efflux, and extracellular chelation (Hou et al. 2020). In other cases, the physical and chemical properties of soil can trap contaminants, making them unavailable to volatilize, leach into water, or be taken up by a plant. As alluded to in Chapter 3, the degradation of contaminants is one of many soil-derived Nature’s Contributions to People.

Soil health is key to this contribution. As reviewed in the case studies, soil organic amendments can increase the absorption of metals and organic chemical contaminants. Changes to pH can reduce the bioavailability of some contaminants. Conversely, contamination can overwhelm the capacity of soil to mitigate risk. High levels of lead and cadmium in soils reduce microbial activity and plant biomass. Microplastics can change soil structure and thereby affect water-holding capacity. The enrichment of pathogens and antibiotic resistance genes in soil microbial communities influenced by microplastics may have implications for human health, but this possibility has not yet been explored.

It is important to note that the contaminants in the case studies presented above were discussed in isolation from one another. Each contaminant class is diverse with heterogeneous impacts based on the nature and quantity of the material and the characteristics of the soil in question. The reality of soil contamination is even more complex because of the potential for co-contamination with multiple compounds. It is likely that the examples of contaminants provided here, along with contaminants found in agricultural inputs such as manure and synthetic fertilizers and pesticides, interact with one another in ways that compound the adverse effects on soil health.

Furthermore, these interacting contaminants are affected by global change factors, such as water stress and high temperatures (Rillig et al. 2019c). There is little available evidence about the effects of many combinations of soil contaminants, in part because it is prohibitively difficult to systematically test the combinatorial effects of large numbers of anything, including contaminants and stressors affecting soils. Given the practical impossibility of testing large numbers of specific combinations, Yang et al. (2022) tested a range of numbers of soil stressors under experimental reactor conditions, drawing at random from a roster of stressors that included various soil contaminants. They found that the positive ecosystem functions of soil microbial diversity were systematically compromised by the action of multiple environmental stressors, including heavy metals, pesticides, plastic film residues, and surfactants as well as nitrogen deposition, heat, drought, salinity, and compaction.

Soil plays a vital role in the cycling of nutrients, where products that are waste to humans, such as manure and biosolids, can be used as sources of energy and nutrients for soil biota and plants. However, soil can no longer be a receptacle for the untreated waste products of agricultural and industrial processes. To incorporate considerations about soil health into decisions about material processing and disposal, the committee suggests that action be taken in the following areas.

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