Next Article in Journal
Insights into the Impact of Extreme Flood Events on Aflatoxin Contamination in Maize from Thessaly, Greece
Previous Article in Journal
Nano- and Microplastics in Single-Use Plastic Water Bottles: A Review of Occurrence, Health Risks, and Regulatory Needs
 
 
Search for Articles:
Title / Keyword
Author / Affiliation / Email
Journal
Article Type
 
 
Advanced Search
 
Section
Special Issue
Volume
Issue
Number
Page
 
Journals
Pollutants
Volume 6
Issue 1
10.3390/pollutants6010016
pollutants-logo
Submit to this Journal Review for this Journal Propose a Special Issue
► Article Menu

Article Menu

share Share announcement Help format_quote Cite question_answer Discuss in SciProfiles
first_page
settings
Order Article Reprints
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:
Background:
Article

Heavy Metal and Nitrate Mobility in Runoff and Seepage Water from a Field Amended with Biochar and Animal Manure

by
George F. Antonious
1,*,
Basanta Neupane
1,
Edward K. Bordoh
1 and
Mohammad H. Dawood
2
1
School of Agriculture and Natural Resources, College of Agriculture, Health, and Natural Resources, Kentucky State University, Frankfort, KY 40601, USA
2
Department of Crop Science, College of Agriculture, University of Telafer, Mosul 41016, Iraq
*
Author to whom correspondence should be addressed.
Pollutants 2026, 6(1), 16; https://doi.org/10.3390/pollutants6010016
Submission received: 30 November 2025 / Revised: 1 February 2026 / Accepted: 25 February 2026 / Published: 3 March 2026
Browse Figures
Versions Notes

Abstract

The undegradable characteristics of heavy metals on environmental quality have become a serious human health concern. A study was conducted in a potato field to investigate the impact of soil amended with animal manure or biochar on the transport of toxic heavy metals and nitrates to runoff and seepage water. The soil in 18 field plots was separated, and each of 3 plots was mixed with biochar, chicken manure, vermicompost, sewage sludge, or cow manure, with 3 plots used as the control. Following a natural rainfall event, the impact of soil treatments on the runoff and infiltration water volume was monitored. Runoff water from the soil amended with biochar exhibited 10.6 L plot−1, whereas cow manure exhibited 4.1 L plot−1, indicating about 61% reduction in runoff water volume. The vermicompost-amended soil increased the seepage water volume from 1.6 L plot−1 in the control treatment to 4.4 L plot−1, indicating a 175% increase in percolating water, a desirable attribute to direct rainfall water towards the plant roots. The concentrations of Pb, Cd, Ni, Mn, Cr, Mg, Cu, and K in infiltration water were greater in runoff sediments, highlighting the need for runoff sediment remediation technology.

Graphical Abstract

1. Introduction

Heavy metals and pesticides are at the top of the list of environmental toxins endangering nature [1]. Heavy metals arise from many sources, such as industry, mining, and agriculture. In the agricultural sector, sources can be categorized into fertilization, pesticides, livestock manure, and wastewater [2]. Xiao et al. [3] have reported that agriculture and industry significantly influence the heavy metal pollution of soil and plants in areas near cement and electroplating factories. Heavy metals’ mobility from agricultural soil to natural water resources has received little attention from soil scientists. Runoff and seepage water together present a fundamental hydrological pathway; runoff carries nutrients and contaminants from the soil rapidly to surface waters, affecting algal blooms and eutrophication, while seepage water slowly leaches soil nutrients and contaminants into groundwater, causing long-term drinking water risk in rural communities when wells are used as a sole source of drinking water. There has been very little research toward a crop-related understanding of amendments’ influence on contaminant mobility, mostly done in generic soil columns or with non-specific plants. Potato (Solanum tuberosum) is a row crop of global importance, commonly cultivated on sloped land with high nitrogen and irrigation input. The ridge and furrow topography, soil disturbance pattern, and different growth phases create a specific soil environment that enhances the risks of runoff and leaching when compared with bare soil [4]. Understanding the influence of amendments in this agroecosystem is therefore important to develop recommendations that are appropriate for the environment and practical farming.
Manure refers to organic materials derived from animal dung, such as cattle and chicken waste, as well as other bulky natural substances, which are applied to the soil as fertilizers with the intention of increasing crop productivity. The nutrient content in manure varies depending on the source, type of feeding, moisture content, storage conditions, bedding and handling methods, and the amount of urine concentrated with the manure.
Animal manure applications in agricultural production have enhanced crop yield and fruit quality at low costs to limited-resource farmers [5,6,7]. However, animal manure is one of the pathways by which heavy metals can enter the ecosystem and accumulate in the soil and discharge into runoff and groundwater sources, absorbed by the growing crops, and consumed by humans and animals, posing a threat to human and animal health [8,9]. Heavy metals influence the microbial population that secretes soil enzymes [10] and may decrease bacterial species richness and both the biomass and diversity of bacterial communities in polluted soils [11]. Zerrari, et al. [12] have reported that Cd is more hazardous to enzymes than Pb, and that Cu suppresses the activity of β-glucosidase more than cellulase activity due to its increased mobility and lesser affinity for soil colloids. Garbowski, et al. [13] have reported that more than 6 Mg ha−1 of world arable soil resources are lost annually due to erosion, which significantly decreases crop yield. Soil amendments could improve the properties of soil, increase crop yields by improving soil fertility and enhancing the availability of nutrients and water for plant uptake, and maintain high microbiological activity in the soil. The benefits of using sewage sludge as a soil amendment are due to its high organic matter content (about 40%), N (4%), and P (2%) [14]. The practice of enriching cattle dung with urea fertilizer as a N source has been found to be effective in overcoming this manure’s common problem of delayed mineralization and low nutrient supply. Due to the rapid growth of the chicken farm industry, chicken manure has become available in large amounts that necessitate a safe disposal method. Overall, manures are a source of almost all of the essential nutrients [15] required for plant growth and production. Vermicompost (cast worms) is considered an outstanding, eco-friendly organic fertilizer when compared with synthetic inorganic mineral fertilizers. Vermicompost increases crop production by reducing pest attacks without contaminating the environment. Agricultural use of vermicompost increases seed germination, stem height, number of leaves, leaf area, leaf dry weight, root length, root number, total yield, number of fruits, leaf chlorophyll content, pH of juice, TSS of juice, micro and macro nutrients, % carbohydrate, and % protein [15].
Biochar from biomass is manufactured for C-sequestration [16], with additional benefits when amended into agricultural that include increasing crop yield and fruit quality due to its higher cation exchange capacity (CEC) when compared with the native soil, which in turn improves soil nutrients retention [17]. The addition of biochar to agricultural soil improves soil porosity and soil alkalinity, and, due to its organic functional groups, can stabilize heavy metals in contaminated soil [18]. Rondon et al. [19] have verified the potential of biochar applications for improving N input in agricultural systems, while indicating the need for long-term field studies to better understand the effect of biochar on biological N2 fixation. When biochar was used in column leaching experiments under controlled conditions to assess its ability to hold nutrients, results indicated that biochar effectively reduced the total amount of NO3, NH4, and PO4 in leachates by 34, 35, and 21%, respectively, relative to native soil [20]. Biochar adsorption of NH3 decreases NH4+ and NO3− ions loss during composting and after manure applications and provides a mechanism for developing slow-release fertilizers [21]. Biochar is currently applied in agricultural production systems to improve soil physical and chemical properties, though the literature review lacks a consideration of the potential risks, induced by the mobility of heavy metals in biochar particles, to natural rainfall when biochar is amended with soil [22]. Zhao, et al. [23] found that biochar released more Pb, Cu, and Zn in runoff water through the soil–biochar matrix.
Consequently, this study aimed to deliver a comprehensive evaluation of the effects of prevalent organic amendments and biochar on pollutant mobility in a potato agroecosystem. We looked at the interrelated system of water partitioning, pollutant mobility, and crop response instead of just looking at specific characteristics in separate reports. The goals were to (1) measure how animal manure (chicken manure (CM), vermicompost (Vermi), municipal sewage sludge (SS), cow dung (Cow)) and biochar affect a natural rainfall event when split into runoff and infiltration paths in a potato production system; (2) evaluate the simultaneous mobility of heavy metals (Cd, Pb, Ni, Zn, Cu, Cr, Mn) and nitrogen species (NH3+ and NO3−) from amended soil into these specific hydrological channels; and (3) assess the comprehensive fate of nitrogen by examining the impact of soil amendment on nitrogen leaching potential and nitrate accumulation in potato tubers, thereby linking environmental loss to crop absorption and food safety.

2. Materials and Methods

2.1. Field Experimental Design

The experiment was conducted on a large scale with a total of 18 plots comprising six treatments with three replicates each that were laid out in randomized complete block design. The study site is situated in an agricultural area with no direct industrial influence. Each experimental plot was 3.7 m wide and 22 m long. The six soil management treatments tested were as follows: cow manure from Black Gold Corporate Company (Oxford, FL, USA), chicken manure from Espoma Company (Millville, NJ, USA), vermicompost from Wiggle Worm Soil Builder (UNCO Industries, Inc., Union Grove, WI, USA), municipal sewage sludge from the Metropolitan Sewer District (Louisville, KY, USA), biochar at the rate of 10% (w/w) from Wakefield Agricultural Carbon, LLC (Valdosta, GA, USA), and an unamended control. All treatments were replicated three times. The control plots consisted of native soil classified as Bluegrass–Maury silty loam, which contains 56% silt, 38% clay, and 6% sand. Application rates of soil amendments, along with their physicochemical properties, are presented in Table 1A,B. Commercially available amendments were used to ensure reproducibility and consistent composition.

2.2. Cultivation Practices

Soil amendments were incorporated into the native soil at 5% N to minimize variations caused by differences in N composition among soil amendments (Table 1A). Each amendment was mixed into the top 15 cm of soil to create standard incorporation practices for potato bed preparation. White potato (Solanum tuberosum, var. Kennebec) was obtained from the Frankfort Service in Kentucky, USA. Potato was selected as the model crop due to its high value and because it is an intensively managed crop that is often grown on sloped land, making it a relevant system for studying runoff and leaching. The Kennebec variety is widely adapted and has a determinate growth habit, providing consistent plant architecture for the experiment. The tubers were cut into appropriately sized pieces, ensuring at least two viable ‘eyes’ per piece, and planted in the prepared field at a depth of 4 inches, with a spacing of 30 cm between rows and plants. A drip irrigation system provided a consistent water supply and weed control measures were implemented as necessary. To prevent weed emergence, the pre-emergence herbicide trifluralin 4EC (Drexel Company, Memphis, TN, USA) was applied at a rate of 0.27 L hectare−1 [24] prior to planting. This herbicide was selected due to its powerful soil adsorption (Koc ~8000), minimizing its potential leaching and runoff.

2.3. Runoff and Seepage Water Collection

Runoff water was collected and measured under natural rainfall events at the lower ends of all plots throughout the growing season, with gutters attached to a tipping bucket apparatus down the slope of the field (Figure 1). From each plot, water that had percolated through the vadose zone was collected by pan lysimeters (4 square feet each) excavated 2 m deep down the soil column while leaving the soil column above it intact. This system facilitated the collection of infiltration water from the vadose zone (the unsaturated water layer that represents the very first entry to groundwater) under normal field conditions (zero tension). Tipping buckets were used to measure the amount of runoff collected down the field slope from each soil treatment. During rainfall, as one bucket filled with water, it tipped down and emptied itself, while the other bucket immediately took its place. Each tipping action is counted in a small switch (counter) which records the exact number of tips and thereby correlates with the amount of runoff for each rainfall event from each soil treatment. Runoff samples were filtered using filter paper no. 1, vacuum-filtered through 0.45 µm cellulose acetate within 2 h; dissolved and particulate metal fractions were extracted and quantified separately using an ICP-MS.

2.4. Heavy Metals and Micronutrients Analyses

During the harvest time, 10 tubers of similar size were randomly selected from each plot. A total of 30 tubers per treatment across three replicates were taken. The tubers were thoroughly washed with tap water, then deionized water, cut into small cubes using a sharp knife, and then oven-dried at 65 °C for 48 h [24]. The dried samples were manually ground using a ceramic mortar and pestle to pass through a 1 mm non-metal sieve. Samples were re-dried in an oven until a constant weight was achieved. One g of each dried sample was mixed with 10 mL of concentrated nitric acid for sample digestion. The mixture was left to stand overnight and then heated for 4 h at 125 °C on a hot plate. The resulting mixture was diluted to 50 mL with double-distilled water and filtered using no. 1 filter paper. The concentrations of Cd, Pb, Cu, Cr, Ni, Zn, Mn, K, and Mg were determined using an inductively coupled plasma–optical emission spectrometer (ICP-OES) (Model 720-ES, Perkin Elmer, Waltham, MA, USA) in standard mode, following USA EPA Method 6010D (ICP-OES) [25]. The detected metal levels in potato tubers were compared against the permissible limits established by the Codex Alimentarius Commission of the Joint FAO/WHO Food Standards [26] and the acceptable limits set by the USA FDA [27].

2.5. Analysis of Potato Ammonia and Nitrates

The potato tubers were collected randomly with five tubers from each replicate. Thus, five fresh potato tubers of comparable size were randomly collected from each replicate (15 tubers from each treatment) and washed with deionized water for chemical analysis. Tubers were cut vertically using a sharpened knife into four quarters, one quarter from each tuber was cut into small cubes and a representative 100 g were selected for sample analysis. Samples were chopped using a kitchen shopper, extracted using 80% ethanol, and filtered using Whatman no. 1 filter paper. Quantification of NH3 and NO3 was carried out using a Fisher brand XL500 Benchtop Meter (Thermo Fisher Scientific, Inc., Rajah Crescent, Singapore, Singapore) equipped with Orion high-performance ammonia and nitrate electrodes (Fisher brand XL500 Benchtop Orion high-performance ammonia and nitrate electrodes) using the methods described by Lipps, et al. [28].

2.6. Statistical Analysis

Statistical analyses of all data obtained from this investigation were performed using SAS software (SAS Institute Inc., Cary, NC, USA, Version 9.1.3) [29]. Wherever ANOVA indicated significant differences (p ≤ 0.05), mean separation was carried out using Duncan’s multiple range test. The one-way analysis of variance, with respect to the effect of amendments on the mobility of heavy metals from soil to runoff, seepage, and sediments, as well as NH3 and NO3 levels in runoff, and accumulation of NO3 in potato tubers, was used in determining the effect of the different soil amendments. The normality and homogeneity of variance assumptions of ANOVA were checked using Shapiro–Wilk and Levene’s tests, respectively. Values are presented as mean ± standard error (SE).

3. Results and Discussion

3.1. Rainfall and Heavy Metals in Runoff and Seepage Water

Figure 2A,B represent the volume of runoff and infiltration water, respectively, collected down the potato field slope of plants grown under six soil management practices. Following a natural rainfall event on 21 July 2023, the runoff water volume from biochar-amended soil was significantly greater (10.64 L plot−1) compared with CM, Vermi, SS, control, and Cow treatments (8.22, 4.6, 4.59, 4.62, 4.28, and 4.13 L Plot−1, respectively) (Figure 2A). Whereas infiltration water volume collected from soil amended with Vermi (4.4 L Plot−1) was greater compared with biochar and the other soil treatments, including the control (Figure 2B). Soil water holding capacity (WHC) is the ability of soil to retain water during rainfall or irrigation events, allowing water to percolate to plant roots. On the other hand, soil or soil mixed with organic amendments has hydrophobicity that repels water instead of absorbing it. Direct and indirect interactions between soil and organic amendments affect WHC. In fact, the hydrophobic nature of some amendments, like biochar, depends on functional groups present on the surface of biochar particles and the original biomass chemical composition. Gelardi and Parikh [30] have reported that not all biochar types are equally efficient in providing the required benefits, due to the variations among the biomass types used in the pyrolysis process, climatic patterns, cropping systems, and soil types. Some plant biomass-derived biochar, such as biochar from rice husk and miscanthus (silver grass) straw, were hydrophobic, whereas other biochar from oilseed rape, softwood, and wheat straw was hydrophilic [31]. The hydrophobic nature of biochar depends on functional groups present on the surface of biochar particles and the feedstock’s chemical composition and pyrolysis temperature. Pyrolyzed biochar appears to be hydrophobic at low temperatures due to the existence of aliphatic compounds, whereas pyrolyzed biochar shows hydrophilicity, as the aliphatic compounds volatilize during the process of hydrolyzation [32,33]. As shown earlier, biochar significantly increased runoff water volume and consequently reduced infiltration (seepage) water in the soil column (Figure 2). This could be due to soil compaction after the addition of biochar to native soil. When biochar is mixed with certain soil types that are rich in clay, such as the Bluegrass silty loam soil in Kentucky (38% clay), compaction reduces the ability of soil to absorb water, causing more runoff due to the poor WHC of clay.
Hydrophilic functional groups (OH, C=O, and COOH) on biochar particles enhance WHC, whereas hydrophobic functional groups (-CH3, CH2 CL2, and C2H4) on biochar particles reduce WHC. However, researchers have reported that the influence of biochar and soil properties on WHC and hydrophobicity remains to be fully understood [31]. On the other hand, the application of organic amendments such as animal manure or biochar significantly impacts soil bulk density. McGrath and Henry [34] have shown that the addition of organic amendment reduced soil compaction and increased soil microporosity and water transport through the soil column (seepage water). This might be due to the role of Vermi in increasing the water percolation in the soil column towards the vadose zone, a desired property that directs the water molecules to the plant roots. Studies have shown that soil compaction can be alleviated by incorporating organic amendments into soils [35] to reduce soil bulk density and facilitate plant root penetration after transplanting.
The results in Figure 3A–C reveal that Vermi and SS treatments significantly (p ≤ 0.05) increased the concentrations of Pb, Cd, and Ni in runoff water compared with the control (no amendment treatment). Pb concentration in runoff water collected from Vermi and SS (97.4 and 96.7 µg plot−1, respectively) was greater compared with 62 µg plot−1 in the control and other animal manure treatments. Similarly, Cd in runoff water increased from 70.5 in the control to 114.5 and 110.8 µg plot−1 in runoff water from Vermi and SS treatments, respectively. Ni also increased in runoff water from 61.6 in the control treatment to 99.0 and 97.7 µg plot−1 in runoff water collected from Vermi and SS treatments, respectively. These results indicate that Pb, Cd, and Ni are released faster from Vermi and SS soil amendments to runoff than from Cow and CM soil amendments. Table 1B shows that Pb concentrations in SS and Vermi are significantly (p ≤ 0.05) lower compared with Cow and CM amendments. However, Pb concentrations in runoff were found to be greater compared with CM and Cow manures (Figure 3A). This could be due to the type and form of Pb present in animal manure. Pb (II) nitrate, Pb (II) acetate, and Pb (II) chloride are highly soluble in water, whereas Pb (II) sulfate and Pb (II) carbonate are generally insoluble in water.
A comprehensive study done by Guo, et al. [36] investigated the toxic effect of Cd on paddy soil properties and reported that Cd caused inhibitory effects on soil microbial activities, microbial growth, and microbial metabolic processes. In addition, ref. [37] confirmed that Cd is highly mobile in the soil, which consequently results in a high toxicity that affects the essential microorganisms and inhibits microbial activities. Usually, soil pH and the content of organic matter in animal manures are the major factors affected by Cd accumulation. In addition, the effects of Pb on soil are several, and include a reduction of soil nutrients, microbial diversity, and soil fertility [37]. Furthermore, earthworms (Eisenia fetida) are usually affected by Pb toxicity, which may cause earthworm mortality. Khan, et al. [38] have revealed the reduced activity of some bacteria and actinomycetes due to the combined toxic effects of Pb and Cd, and have reported that soil pH, organic matter, ionic exchange capacity, and texture affect Pb accumulation in the soil. Figure 3C reveals that Ni concentrations in runoff water were found to be low in CM, Cow, and the control treatments compared with Vermi and SS treatments. Heavy metals can be present in agricultural soil in both solid and solution phases. In the solid phase, they are immobile, inert and harmless because they are immobilized through adsorption of organic and inorganic components of the soil particles or may be precipitated as pure solids. Whereas in the aqueous phase heavy metals are mobile and toxic. Metal ions in the solid phase may become available if there is change in soil condition, pH or oxidation-reduction potential. Metals can also exist in the soil solution phase as free medallions or as soluble complexes with inorganic (OH−, CO32−, HCO3−, SO42−, NO3−, and Cl−) or organic ligands [39,40].
The results in Figure 4 provide critical insights into the chemical composition of soil amendments and their potential environmental and agronomic impacts. These results also revealed that the concentrations of Mn, Cr, and Mg in runoff water from the Vermi treatment were significantly greater compared to the control treatment. Mn was generally greater in runoff water from animal manure compared to the control treatment (Figure 4A). While Mn is an essential micronutrient for plants, elevated levels can become toxic, leading to issues like nutrient lockout or Mn toxicity in plants, especially in acidic soils. The substantial increase in total Cr in runoff water from Vermi and SS application (Figure 4B) increases concerns about soil toxicity, potential groundwater contamination, and the long-term health of the soil food web. The figure strongly suggests that while soil amendments are effective at supplying bulk nutrients and metals, it carries a significant risk of heavy metal pollution (specifically Mn and Cr) in runoff following natural rainfall events.
Figure 5A indicates that the concentration of Cu in runoff water from biochar-amended soil was significantly (p ≤ 0.05) reduced to 66.9 µg plot−1 compared with SS and Vermi treatments (95 and 82 µg plot−1, respectively). K and Zn in runoff from biochar-amended soils were also reduced compared with Vermi-amended soil (Figure 5B,C). In the past decade, biochar initiation from numerous feedstocks has been investigated for the remediation of metals/metalloids contaminated soils. Most of this work has focused on cationic metals such as Cd, Cu, Pb, Ni, and Zn. Recently, results have indicated that Biochar can restore degraded soil and can play a significant role in preventing water molecules and nutrient minerals in the upper soil surface from evaporating following rainfall or rainfall and irrigation events, as it contains numerous tiny tunnels that retain water and minerals in the soil layers, increasing water availability for plants. This boosts crop cost-savings and decreases the need for inorganic elemental fertilizers [5]. Biochar’s application to soil increases soil fertility by improving nutrient retention and uptake by developing plants. In addition, biochar adsorbs nutrients on its surfaces, reducing the amounts of nutrients (NO3−–N, NH4+–N, and PO43−–P) that leach after animal manure and inorganic fertilizer application [41]. Inorganic-N (NO3−–N, NH4+–N) is adsorbed on the surface of biochar and is released gradually in the soil solution, thus functioning as a slow-release fertilizer [42].
Several mechanisms of immobilization appear to be responsible for the reduction in the metal concentration in the runoff being carried by biochar-amended soils. Thus, the negatively charged surfaces of biochar hold major metal ions and prevent their movement and availability to biological activity. Biochar, being highly alkaline, will also tend to form insoluble hydroxides or carbonates of metals by making such metals precipitate out. Interestingly, functional groups present on the surface of biochar, such as carboxyl, hydroxyl, and phenolic moieties, enhance binding with heavy metals via complex reactions. It is exactly these interactions that together could cause a decrease in copper, potassium, and zinc concentrations in runoff from biochar-treated plots and which thus underline its potential for reducing the transport of metals in agricultural systems [43].

3.2. Heavy Metals in Seepage Water

Figure 6 represents the concentration of Pb, Cd, and Ni in seepage water collected after the application of animal manure and biochar to a potato field following a natural rainfall event.
Lead (Pb) concentration was significantly (p ≤ 0.05) greater (0.85 µg plot−1) in Vermi-amended seepage water when compared with seepage water from CM, Cow, unamended soil (control treatment), and biochar. Pb concentration in seepage water from CM (0.27 µg plot−1) treatment was significantly lower compared with other treatments (Figure 6A). This trend of high Pb concentration in seepage water from Vermi was also observed in runoff water collected from Vermi-amended soil. As described earlier, Vermi has a more neutral pH compared with some animal manures, which can affect the solubility and mobility of Pb soluble forms (lead nitrate [Pb (NO3)2], lead acetate (PbCOOCH3) and lead chloride (PbCl2) through the soil column. In fact, hazardous materials associated with a heavy metal-contaminated location depend on the metal’s chemical form. Sites in which metals exist mainly in residual forms (strongly bound within the crystal lattice), like lead sulfide, are difficult to extract and have little or no hazard to the environment and living organisms. For example, Pb in the form of lead phosphates (Pb3 (PO4)2 and PbHPO4), mixed lead chloride phosphate (Pb5(PO4)3Cl), and PbSO4 exist mainly in the residual form, where they present little or no risks to the environment [44]. On the other hand, lead nitrate [Pb (NO3)2], lead acetate (PbCOOCH3), and lead chloride (PbCl2) are in the form of non-residual heavy metals that are readily released and soluble in soil, and have a high potential for mobility and toxicity. Pb in these forms may be taken up and accumulated by living organisms or be leached down the soil profile, reaching the groundwater system [45].
This observation was also similar that of Cd and Ni concentrations in seepage water collected from Vermi-amended soil (Figure 6B,C). Cd occurs naturally with Zn and Pb in sulfide ores. The elevated concentrations of Cd in air, water, and soil occur in areas close to industrial emission sources, particularly those of mining and metal processing operations. Cd is a highly toxic metal that can accumulate in the human body and cause irreversible damage to a number of biological systems, even at an extremely low dose. Figure 6C reveals that Ni concentrations in seepage water collected from Vermi and SS were greater than CM, Cow, biochar and the control treatments. Soil amended with biochar reduced the concentration of Ni from 0.7 and 0.6 µg plot−1 from Vermi and SS, respectively, to 0.41 µg plot−1. This represents about a 43% reduction in Ni concentration in seepage water due to the application of biochar. This also indicates that SS and Vermi are highly porous amendments that facilitate the mobility of Ni from soil into the vadose zone (the unsaturated water layer below the plant root), the first entry to the groundwater layer. Exposure to Ni in water at concentrations above the recommended limits of 0.10 µg mL−1 can have adverse effects on human health. The most common adverse health effects of Ni in humans are allergic reactions, and a high concentration of Ni can cause lung and nasal sinus cancers. For comparison purposes, the concentrations of Ni and other metals in seepage water collected from the vadose zone were calculated and expressed as µg plot−1. Accordingly, transferring these concentrations to µg mL−1, as shown in Table 2B, revealed that the detected concentrations of Pb, Cd, Ni and other heavy metals in seepage water collected from the vadose zone were all below the allowable limits of 0.1 µ mL−1 established by the US EPA [46].
Figure 7A,B reveal that Cr and Cu concentrations in seepage water collected from Vermi and SS amended soil were greater compared with all other amendments, including the control treatment. This could be due to the high permeability and pore size of SS and the Vermi particle size. There were no significant differences between biochar and the control treatments in the mobility of Cr and Cu from soil to seepage water, in spite of the high surface area of biochar particles. Mg is usually washed from soil rocks and subsequently ends up in water bodies. Chemical industries add Mg plastics and other materials as a fire protection measure or as a filler and it also ends up in the environment from fertilizer application and from cattle feed. Mg concentrations (Figure 7C) in seepage water were higher in the SS and biochar treatments and lowest in the CM treatment. Mg is an essential element that contributes to chlorophyll production and metal stress tolerance [47].
Copper (Cu) is essential for healthy growth. However, high doses of Cu salts to can be extremely harmful to humans and animals. Cu can enter the human body through dust, food, and water. Free Cu ion (Cu2+) is one of the most toxic forms of Cu in aquatic life. Cr is a transition metal that is extensively used in various industries. On a worldwide basis, around 80% of Cr is employed in metallurgical applications. It is added as a biocide in cooling water to prevent corrosion. Consumption of Cu-contaminated water or foods can cause acute gastrointestinal symptoms [50,51]. An intake of high amounts of Cu salt can cause nausea and acute gastric irritation. However, the present investigation indicated that Cr and Cu concentrations in seepage water collected from the six soil treatments expressed as µg mL−1 were below the allowable limits (Table 2). On the contrary, runoff water was contaminated with the three heavy metals, Pb, Cd, and Mn, that were all above the drinking water allowable limits, as shown in Table 2A. However, a significant portion of the runoff water volume and metal and nutrient concentrations in runoff will be intercepted by the majority of water in the rivers, ponds, and streams down the field slope upon reaching the natural water resources, where it will be diluted by cleaner waters.
Figure 8A represents the concentrations of K in infiltration water from six soil management practices. The concentration of K was greatest (189 µg plot−1) in seepage water collected from Vermi treatments compared with other treatments including the biochar. This concentration is 0.21 µg mL−1, which is far below the allowable limit of 12 µg L−1 in drinking water (Table 2B). K is necessary for organisms and can be found in all humans, animals, and plant cells. K is vital for human body functions like heart protection, regulation of blood pressure, protein dissolution, muscle contraction, nerve stimulus etc. Zn and Mn concentrations in seepage water fluctuated among the water collected from the vadose zone (Figure 8B). Zn concentration was greatest in seepage water collected from the Vemi (3 µg plot−1) treatment and was detected at very low concentration (0.32 µg plot−1) in seepage water from SS treatments. These differences could be due to the variability of the organic matter and the pore space size of the manure, which control the mobility through the soil column into the vadose zone. Zn is considered an essential micronutrient and a much less toxic metal. However, excessive amounts of Zn intake (e.g., inhaling Zn vapors and ingesting Zn-contaminated food or water) can induce system dysfunctions, which cause an impairment of growth and reproduction [50]. Zn is released into the environment mainly from anthropogenic activities such as mining, steel production, coal burning, and burning of waste. Unlike other heavy metals such as Pb and Hg, Zn contamination is generally not highly concerned.
Seepage water in wells in some counties in Kentucky is used as a sole source of drinking water. Mn concentrations in seepage water collected from the control treatment (non-amended soil) were highly significant (p ≤ 0.05) compared with CM and biochar treatments (Figure 8C). These results can be explained by the high concentration of Mn in native soil (1262.5 µg g−1 soil) compared with the concentration of Mn in all other soil treatments, as shown in Table 1B. Mn in drinking water can cause visual and operational problems. Potential problems occur when dissolved Mn (II) is oxidized to insoluble forms, such as, Mn (III) and Mn (IV). These particles can impart turbidity and cause a black-brown color to drinking water. The US EPA conducted a health-based assessment of Mn and established a lifetime health advisory level of 0.05 µg mL−1 for Mn in drinking water [46]. Anthropogenic activities such as fertilizer application, soil amendment addition, pesticide application, animal waste used as organic fertilizers, auto emissions, and biochar application to agricultural soils can introduce greater amounts of heavy metals to soils that contaminate natural water resources [51,52].
When concentrations of Mn exceed the notification level, certain requirements and recommendations apply to all public water systems, which would require some utilities to change their Mn management approach [53].
Metals are usually distributed throughout the soil particle mechanism in various geochemical forms (Figure 9) including ion exchange, adsorption, precipitation, and complexation. Due to changes in environmental conditions, such as acidification by acid rain, redox potential or organic ligand concentrations, oxidation and other things can cause mobilization of metals from solid to liquid phase. Mobility and adsorption of metals accumulated in sediments can cause contamination of surrounding waters and represent a potential risk for an aquatic environment.
Geochemical reactions, including adsorption and desorption, precipitation and dissolution, reduction and oxidation are governing processes of the bioavailability and toxicity of reactive metal in soils and from soil to runoff water and sediment. Exchange between aqueous solutions and particle sites with static electric charges on the surfaces of minerals is an important reaction influencing the reaction of ionic particles in soils. Due to the negative charge borne by soil colloids, cation exchange reactions can heavily influence the retention and transport of toxic metal elements. Rainfall events generate runoff that removes soluble metals from agricultural fields to surface water bodies.
The pH of each amendment (Table 1B) significantly influenced metal mobility. The two amendments with lower pH values, sewage sludge at pH 5.72 and vermicompost at pH 6.57, enhanced the solubility and mobility of lead (Pb), cadmium (Cd), and nickel (Ni) in runoff water. The two amendments with higher pH values, biochar at pH 7.85 and cow manure at pH 8.09, increased metal adsorption and precipitation, which resulted in lower metal levels in the seepage water towards the vadose zone. The pH-dependent behavior of the system illustrates the geochemical mechanisms depicted in Figure 9, which elucidate how soil pH regulates ion exchange, adsorption, and complexation processes.
Sediment contamination by heavy metals could be used as an important index for monitoring the degree of natural water resources. There is no single universal maximum residue limits of heavy metals in sediment as a transported residue of the soil. However, the mobility of the non-residual fraction of heavy metals (metals that are not tightly bound, mobile, and bioavailable) obtained from the present investigation expressed as µg g−1 sediment can be categorized as follows: K > Zn > Mg > Cd > Ni = Cu > Pb > Cr.

3.3. Nitrates and Ammonia in Runoff, Seepage Water, and Potato Tubers

Figure 10A reveals that the concentration of NH3+–N was found to be greatest in the Vermi and SS treatments (289 and 298 µg plot−1, respectively) and lowest in runoff collected from CM soil treatments (187 µg plot−1). Though NO3–N concentrations were greatest in CM compared with other animal manures, biochar, and the control treatments. Figure 10B indicates that the concentrations of NH3+–N and NO3–N in seepage water collected from the biochar-amended soil were significantly greater (63 and 180 µg plot−1, respectively) whereas the unamended soil (control treatment) had the lowest concentrations. Rainfall intensity and duration influence the concentration and overall loss of N in manure and the mobility of N with regard to runoff and groundwater. The relationship between potential loss and N application rate in agricultural production systems is critical to establishing environmentally sound manure management guidelines [53]. The timing of manure applications relative to the occurrence of rainfall influences N loss to runoff and seepage water If manure applications are made during periods of the year when intense storms are likely, meaning that the percentage of applied N loss is higher than if applications of manure are made when runoff probabilities are lower.
NO3− concentration above the allowable limit is a water quality concern for several reasons: it is linked to methemoglobinemia in infants, to toxicities in livestock, and to increased eutrophication in both fresh and saline (estuary) waters. The EPA has established a maximum contaminant level (MCL) for NO3–N in drinking water of 10 mg L−1 to protect babies under 3–6 months of age, for whom bacteria that live in their digestive tract can reduce NO3− to nitrite (NO2−), causing hemoglobin to transform into methemoglobin, which in turn interferes with the oxygen-carrying ability of blood. NO3− can also be toxic to livestock if reduced to NO2− where it can cause methemoglobinemia, which causes abortions in cattle. The tolerance level for livestock is about 40 mg NO3–N L−1, which is higher than for humans. Levels of 40–100 mg NO3–N L−1 in drinking water are considered risky.
Figure 11 represents the concentrations of NH3+–N and NO3–N in potato tubers of plants grown under six soil management practices. Generally, vegetables contain NO3− at varying levels, ranging from 1 to 10,000 mg kg−1 [54]. Vegetables can be classified according to their NO3− content into very low (<200), low (200 to <500), middle (500 to <1000), high (1000 to <2500), and very high, (>2500 mg 100−1 g fresh weight) [55]. In fact, the primary variables for NO3− human intake include the type of vegetables consumed, NO3− levels in the vegetable type, and the amounts of vegetables consumed daily. The mean total NO3− daily intake per person in Europe ranges between 50 and 140 mg and in the USA about 40–100 mg [56]. Toxic doses with methemoglobin formation as a criterion for toxicity ranged from 33 to 350 mg NO3− ion kg−1 BW [57] and human lethal doses of 67–833 mg NO3− ion kg−1 body weight (BW) have been reported. Consumption of one serving of a NO3−–rich food or supplement can exceed the World Health Organization’s acceptable daily intake for NO3− (0–3.7 mg kg−1 body weight or 222 mg day−1 for a 60 kg adult).
The greatest concentration was detected in potato tubers of plants grown in CM and Bio treatments compared with plants grown in non-amended soil (the control). The results also reveal those plants grown in Vermi had the lowest concentration of NO3. Compared with the other animal manures, the maximum NO3 concentration ingested daily on fresh weight bases is less than 3.65 mg kg−1 expressed in terms of body weight (BW). Therefore, a person with an average weight of 70 kg should not consume more than 255.5 mg of NO3 daily [58]. Considering a value of 34.89 µg g−1 of NO3 in potato tubers grown in CM-amended soil, consuming 200 g will generate 200 × 34.89 mg NO3 would be equal to 6979.2/70 or 99.7 mg kg−1 BW. Accordingly, none of the investigated treatments represent any NO3 hazard to consumers.

4. Conclusions

Unlike organic pollutants, natural processes of decomposition do not remove heavy metals in soil and water. Heavy metals usually possess significant toxicity to aquatic organisms and human health through the food chain. Therefore, investigating the transformation and removal mechanisms of heavy metal in soil, runoff water, and sediment becomes necessary. A field study was conducted to investigate the impact of animal manure (chicken manure, vermicompost, sewage sludge, cow manure) and biochar on the volume of runoff and seepage water following natural rainfall events, and we monitored the mobility of heavy metals and nitrates in animal manures to runoff water and sediment. The results reveal that the runoff water was greatest in soil amended with biochar, whereas infiltration water was greatest from soil amended with vermicompost compared with all other treatments, including the control. The concentrations of Pb, Cd, Ni, Mn, Cr, Mg, Cu, K, and Zn in the infiltration (seepage) water collected from the vadose zone were lower than the drinking water standards established by the EPA. Meanwhile, heavy metals were greater in runoff sediments when compared with runoff water, highlighting the need for runoff sediment remediation technology.
The study demonstrated that customized soil amendments can be integrated into pilot-scale systems for soil and water treatment. The biochar can function as a metal runoff control agent when applied to edge-of-field buffer strips or constructed wetlands, while vermicompost will enhance water retention. The basic research studies must assess performance over time, financial viability, and ability to expand under different agricultural and climate conditions so that the amendments can be effectively used in sustainable land and water management programs.
Animal manure and biochar can be valuable resources if managed properly using cost-effective BMPs at an adjustable rate and by timing their application by avoiding the incidence of rainfall events that enhance the mobility of heavy metals to natural water resources. However, in areas of intensively confined animal operations, where manure production and application exceed local crop N requirements, the effect of manure on soil N transformations and heavy metal mobility to runoff water and sediments could be reduced. Spokas, et al. [59] have reported that the influence of biochar on soil properties may be positive or negative depending on the quality and rate of biochar application. Some biochar nutrients are leachable despite the observations of nutrient sorption [60]. Our future objectives will include researching the impact of heavy metals in animal manure on soil enzyme activity.

Author Contributions

Writing, original ideas, field design, methodology, review, and editing: G.F.A.; field work, sample collection, laboratory analysis, and data collection: B.N. and E.K.B. Statistical analysis: M.H.D. All authors have read and agreed to the published version of the manuscript.

Funding

This research project was funded by the United States Department of Agriculture, National Institute of Food and Agriculture (USDA/NIFA), under a grant to Kentucky State University, Agreement #KYX-10-23-80P, Accession 7005611.

Data Availability Statement

The original contributions presented in the study are included in the article; further inquiries can be directed to the corresponding author at george.antonious@kysu.edu.

Acknowledgments

The authors wish to express their appreciation for E. Turley, Anjan Nepal, and the farm crew at Kentucky State University Research Farm for maintaining the field experiment.

Conflicts of Interest

The authors declare no conflicts of interest.

References

  1. Alengebawy, A.; Abdelkhalek, S.T.; Qureshi, S.R.; Wang, M.-Q. Heavy Metals and Pesticides Toxicity in Agricultural Soil and Plants: Ecological Risks and Human Health Implications. Toxics 2021, 9, 42. [Google Scholar] [CrossRef]
  2. Li, Z.; Ma, Z.; van der Kuijp, T.J.; Yuan, Z.; Huang, L. A review of soil heavy metal pollution from mines in China: Pollution and health risk assessment. Sci. Total Environ. 2014, 468–469, 843–853. [Google Scholar] [CrossRef]
  3. Xiao, R.; Wang, S.; Li, R.; Wang, J.J.; Zhang, Z. Soil heavy metal contamination and health risks associated with artisanal gold mining in Tongguan, Shaanxi, China. Ecotoxicol. Environ. Saf. 2017, 141, 17–24. [Google Scholar] [CrossRef] [PubMed]
  4. Kumar, V.; Rawat, A.; Rao, D. Population ecology of soybean-rhizobia in diverse crop rotations in Central India. Agric. Ecosyst. Environ. 2017, 240, 261–268. [Google Scholar] [CrossRef]
  5. Antonious, G.F.; Dawood, M.H.; Turley, E.T.; Trivette, T.G. Soil Amendments Enhanced Summer Squash Yield, Fruit Composition, Quality, and Soil Enzymes Activity. Agric. Sci. 2022, 13, 684–701. [Google Scholar] [CrossRef]
  6. Antonious, G.F. Duality of Biochar and Organic Manure Co-Composting on Soil Heavy Metals and Enzymes Activity. Appl. Sci. 2025, 15, 3031. [Google Scholar] [CrossRef]
  7. Antonious, G.F.; Nepal, A.; Neupane, B. Remediation of Cd, Cu, and Zn Metals in Soil Amended with Biochar and Animal Manure Using a Hyperaccumulator. Pollutants 2025, 5, 22. [Google Scholar] [CrossRef]
  8. Hejna, M.; Moscatelli, A.; Onelli, E.; Baldi, A.; Pilu, S.; Rossi, L. Evaluation of concentration of heavy metals in animal rearing system. Ital. J. Anim. Sci. 2019, 18, 1372–1384. [Google Scholar] [CrossRef]
  9. Lv, Y.; Bao, J.; Liu, D.; Gao, X.; Yu, Y.; Zhu, L. Synergistic effects of rice husk biochar and aerobic composting for heavy oil-contaminated soil remediation and microbial community succession evaluation. J. Hazard. Mater. 2023, 448, 130929. [Google Scholar] [CrossRef]
  10. Deng, Y.F.; Yue, X.B.; Cui, Y.J.; Shao, G.H.; Liu, S.Y.; Zhang, D.W. Effect of pore water chemistry on the hydro-mechanical behaviour of Lianyungang soft marine clay. Appl. Clay Sci. 2014, 95, 167–175. [Google Scholar] [CrossRef]
  11. Guo, H.; Luo, S.; Chen, L.; Xiao, X.; Xi, Q.; Wei, W.; Zeng, G.; Liu, C.; Wan, Y.; Chen, J.; et al. Bioremediation of heavy metals by growing hyperaccumulaor endophytic bacterium Bacillus sp. L14. Bioresour. Technol. 2010, 101, 8599–8605. [Google Scholar] [CrossRef] [PubMed]
  12. Zerrari, N.; Rais, N.; El Ghachtouli, N.; Kouchou, A.; Ijjaali, M. Heavy Metals Effects on Agricultural Soil Enzyme Activities of Fez, Morocco. J. Ecol. Eng. 2023, 24, 144–154. [Google Scholar] [CrossRef]
  13. Garbowski, T.; Bar-Michalczyk, D.; Charazińska, S.; Grabowska-Polanowska, B.; Kowalczyk, A.; Lochyński, P. An overview of natural soil amendments in agriculture. Soil Tillage Res. 2023, 225, 105462. [Google Scholar] [CrossRef]
  14. Onunwa, A.O.; Igwe, C.A.; Mbonu, O.A. Tillage and Residual effect of some Organic Amendments on Aggregate associated Soil Carbon of an Ultisol, Growth and Yield of Maize and Cowpea Intercrop. Niger. J. Soil Sci. 2019, 28, 31–41. [Google Scholar] [CrossRef]
  15. Joshi, R.; Singh, J.; Vig, A.P. Vermicompost as an effective organic fertilizer and biocontrol agent: Effect on growth, yield and quality of plants. Rev. Environ. Sci. Bio/Technol. 2014, 14, 137–159. [Google Scholar] [CrossRef]
  16. Yang, X.Y.; Chang, K.-H.; Kim, Y.J.; Zhang, J.; Yoo, G. Effects of different biochar amendments on carbon loss and leachate characterization from an agricultural soil. Chemosphere 2019, 226, 625–635. [Google Scholar] [CrossRef]
  17. Ippolito, J.A.; Laird, D.A.; Busscher, W.J. Environmental Benefits of Biochar. J. Environ. Qual. 2012, 41, 967–972. [Google Scholar] [CrossRef]
  18. Shen, Z. Perspectives of biochar application in soil to immobilize heavy metals. In Biochar Application in Soil to Immobilize Heavy Metals; Elsevier: Amsterdam, The Netherlands, 2024; pp. 239–244. [Google Scholar] [CrossRef]
  19. Rondon, M.A.; Lehmann, J.; Ramírez, J.; Hurtado, M. Biological nitrogen fixation by common beans (Phaseolus vulgaris L.) increases with bio-char additions. Biol. Fertil. Soils 2006, 43, 699–708. [Google Scholar] [CrossRef]
  20. Banik, C.; Koziel, J.A.; De, M.; Bonds, D.; Chen, B.; Singh, A.; Licht, M.A. Biochar-Swine Manure Impact on Soil Nutrients and Carbon Under Controlled Leaching Experiment Using a Midwestern Mollisols. Front. Environ. Sci. 2021, 9, 609621. [Google Scholar] [CrossRef]
  21. Clough, T.; Condron, L.; Kammann, C.; Müller, C. A Review of Biochar and Soil Nitrogen Dynamics. Agronomy 2013, 3, 275–293. [Google Scholar] [CrossRef]
  22. Shen, Z. Long-term stability of heavy metals in biochar-treated soil. In Biochar Application in Soil to Immobilize Heavy Metals; Elsevier: Amsterdam, The Netherlands, 2024; pp. 131–162. [Google Scholar] [CrossRef]
  23. Zhao, L.; Nan, H.; Kan, Y.; Xu, X.; Qiu, H.; Cao, X. Infiltration behavior of heavy metals in runoff through soil amended with biochar as bulking agent. Environ. Pollut. 2019, 254, 113114. [Google Scholar] [CrossRef]
  24. Antonious, G.F.; Snyder, J.C. Accumulation of heavy metals in plants and potential phytoremediation of lead by potato, Solanum tuberosum L. J. Environ. Sci. Health Part A 2007, 42, 811–816. [Google Scholar] [CrossRef]
  25. Leykin, A.; Nelms, S. Inductively Coupled Plasma Mass Spectrometry in Environmental Analysis. In Comprehensive Environmental Mass Spectrometry; ILM Publications: Srinagar, India, 2012; p. 225. [Google Scholar]
  26. Havlik, S. WHO-FAO Codex Alimentarius Commission (CAC). In Encyclopedia of Food Safety; Elsevier: Amsterdam, The Netherlands, 2024; pp. 734–744. [Google Scholar] [CrossRef]
  27. Sidhu, G.P.S.; Singh, H.P.; Batish, D.R.; Kohli, R.K. Effect of lead on oxidative status, antioxidative response and metal accumulation in Coronopus didymus. Plant Physiol. Biochem. 2016, 105, 290–296. [Google Scholar] [CrossRef] [PubMed]
  28. Lipps, W.C.; Braun-Howland, E.B.; Baxter, T.E. Standard Methods for the Examination of Water and Wastewater; APHA Press: Washington, DC, USA, 2023. [Google Scholar]
  29. SAS Institute Inc. SAS 9.4 Formats and Informats: Reference; SAS Institute Inc.: Cary, NC, USA, 2023. [Google Scholar]
  30. Gelardi, D.L.; Parikh, S.J. Soils and Beyond: Optimizing Sustainability Opportunities for Biochar. Sustainability 2021, 13, 10079. [Google Scholar] [CrossRef]
  31. Adhikari, S.; Timms, W.; Mahmud, M.A.P. Optimising water holding capacity and hydrophobicity of biochar for soil amendment—A review. Sci. Total Environ. 2022, 851, 158043. [Google Scholar] [CrossRef] [PubMed]
  32. Zheng, X.; Li, X.; Singh, B.P.; Wei, L.; Huang, L.; Huang, Y.; Huang, Q.; Chen, X.; Su, Y.; Liu, Z.; et al. Biochar protects hydrophilic dissolved organic matter against mineralization and enhances its microbial carbon use efficiency. Sci. Total Environ. 2021, 795, 148793. [Google Scholar] [CrossRef]
  33. Zornoza, R.; Moreno-Barriga, F.; Acosta, J.A.; Muñoz, M.A.; Faz, A. Stability, nutrient availability and hydrophobicity of biochars derived from manure, crop residues, and municipal solid waste for their use as soil amendments. Chemosphere 2016, 144, 122–130. [Google Scholar] [CrossRef]
  34. McGrath, D.; Henry, J. Organic amendments decrease bulk density and improve tree establishment and growth in roadside plantings. Urban For. Urban Green. 2016, 20, 120–127. [Google Scholar] [CrossRef]
  35. Olson, N.C.; Gulliver, J.S.; Nieber, J.L.; Kayhanian, M. Remediation to improve infiltration into compact soils. J. Environ. Manag. 2013, 117, 85–95. [Google Scholar] [CrossRef]
  36. Guo, Y.; Cheng, S.; Fang, H.; Yang, Y.; Li, Y.; Zhou, Y. Responses of soil fungal taxonomic attributes and enzyme activities to copper and cadmium co-contamination in paddy soils. Sci. Total Environ. 2022, 844, 157119. [Google Scholar] [CrossRef]
  37. Dotaniya, M.L.; Dotaniya, C.K.; Solanki, P.; Meena, V.D.; Doutaniya, R.K. Lead Contamination and Its Dynamics in Soil–Plant System. In Radionuclides and Heavy Metals in the Environment; Springer: Cham, Switzerland, 2019; pp. 83–98. [Google Scholar] [CrossRef]
  38. Khan, S.; El-Latif Hesham, A.; Qiao, M.; Rehman, S.; He, J.-Z. Effects of Cd and Pb on soil microbial community structure and activities. Environ. Sci. Pollut. Res. 2009, 17, 288–296. [Google Scholar] [CrossRef] [PubMed]
  39. Agrelli, D.; Caporale, A.G.; Adamo, P. Assessment of the Bioavailability and Speciation of Heavy Metal(loid)s and Hydrocarbons for Risk-Based Soil Remediation. Agronomy 2020, 10, 1440. [Google Scholar] [CrossRef]
  40. Praveena, S.M.; Pradhan, B.; Aris, A.Z. Assessment of bioavailability and human health exposure risk to heavy metals in surface soils (Klang district, Malaysia). Toxin Rev. 2017, 37, 196–205. [Google Scholar] [CrossRef]
  41. Jeyasubramanian, K.; Thangagiri, B.; Sakthivel, A.; Dhaveethu Raja, J.; Seenivasan, S.; Vallinayagam, P.; Madhavan, D.; Malathi Devi, S.; Rathika, B. A complete review on biochar: Production, property, multifaceted applications, interaction mechanism and computational approach. Fuel 2021, 292, 120243. [Google Scholar] [CrossRef]
  42. Haider, G.; Steffens, D.; Moser, G.; Müller, C.; Kammann, C.I. Biochar reduced nitrate leaching and improved soil moisture content without yield improvements in a four-year field study. Agric. Ecosyst. Environ. 2017, 237, 80–94. [Google Scholar] [CrossRef]
  43. Ahmad, M.; Lee, S.S.; Lee, S.E.; Al-Wabel, M.I.; Tsang, D.C.W.; Ok, Y.S. Biochar-induced changes in soil properties affected immobilization/mobilization of metals/metalloids in contaminated soils. J. Soils Sediments 2016, 17, 717–730. [Google Scholar] [CrossRef]
  44. Jin, X.; Tong, X.; Hua, S.; Xu, Y. Ecological and Health Risk Assessment of Soil Heavy Metal Contamination Along National Highway 107 in China. Ecol. Chem. Eng. S 2024, 31, 155–175. [Google Scholar] [CrossRef]
  45. Parvin, A.; Moniruzzaman, M.; Hossain, M.K.; Saha, B.; Parvin, A.; Suchi, P.D.; Hoque, S. Chemical Speciation and Potential Mobility of Heavy Metals in Organic Matter Amended Soil. Appl. Environ. Soil Sci. 2022, 2022, 2028860. [Google Scholar] [CrossRef]
  46. U.S. Environmental Protection Agency. Drinking Water Regulations and Contaminants. 2025. Available online: https://www.epa.gov/sdwa/drinking-water-regulations-and-contaminants (accessed on 12 February 2026).
  47. White, P.J.; Broadley, M.R. Biofortification of crops with seven mineral elements often lacking in human diets—Iron, zinc, copper, calcium, magnesium, selenium and iodine. New Phytol. 2009, 182, 49–84. [Google Scholar] [CrossRef]
  48. McCartney, D. J. Mann &amp; A. S. Truswell (editors). Essentials of Human Nutrition Oxford: Oxford University Press2007 £35.99 (paperback) pp. 599 ISBN 978-0-19-929097-0. Br. J. Nutr. 2008, 99, 1168. [Google Scholar] [CrossRef]
  49. Goncalves, A.C.; Nacke, H.; Schwantes, D.; Coelho, G.F. Heavy Metal Contamination in Brazilian Agricultural Soils due to Application of Fertilizers. In Environmental Risk Assessment of Soil Contamination; Hernandez-Soriano, M.C., Ed.; Intech: London, UK, 2014; Chapter 4; pp. 105–135. [Google Scholar] [CrossRef]
  50. Srivastava, A.; Peshin, S.S.; Kaleekal, T.; Gupta, S.K. An epidemiological study of poisoning cases reported to the National Poisons Information Centre, All India Institute of Medical Sciences, New Delhi. Hum. Exp. Toxicol. 2005, 24, 279–285. [Google Scholar] [CrossRef]
  51. Kozisek, F. Regulations for calcium, magnesium or hardness in drinking water in the European Union member states. Regul. Toxicol. Pharmacol. 2020, 112, 104589. [Google Scholar] [CrossRef] [PubMed]
  52. Urra, J.; Alkorta, I.; Garbisu, C. Potential benefits and risks for soil health derived from the use of organic amendments in agriculture. Agronomy 2019, 9, 542. [Google Scholar] [CrossRef]
  53. Wang, Z.; Zhang, T.Q.; Tan, C.S.; Wang, X.; Taylor, R.A.J.; Qi, Z.M.; Yang, J.W. Modeling the Impacts of Manure on Phosphorus Loss in Surface Runoff and Subsurface Drainage. J. Environ. Qual. 2019, 48, 39–46. [Google Scholar] [CrossRef] [PubMed]
  54. Correia, M.; Barroso, Â.; Barroso, M.F.; Soares, D.; Oliveira, M.B.P.P.; Delerue-Matos, C. Contribution of different vegetable types to exogenous nitrate and nitrite exposure. Food Chem. 2010, 120, 960–966. [Google Scholar] [CrossRef]
  55. Hmelak Gorenjak, A.; Cencič, A. Nitrate in vegetables and their impact on human health. A review. Acta Aliment. 2013, 42, 158–172. [Google Scholar] [CrossRef]
  56. Tamme, T.; Reinik, M.; Roasto, M.; Juhkam, K.; Tenno, T.; Kiis, A. Nitrates and nitrites in vegetables and vegetable-based products and their intakes by the Estonian population. Food Addit. Contam. 2006, 23, 355–361. [Google Scholar] [CrossRef]
  57. FAO; WHO. Compendium of Food Additive Specifications. Joint FAO/WHO Expert Committee on Food Additives (JECFA), 87th Meeting June 2019; FAO JECFA Monographs 23; FAO: Rome, Italy; WHO: Geneva, Switzerland, 2020. [Google Scholar]
  58. Salehzadeh, H.; Maleki, A.; Rezaee, R.; Shahmoradi, B.; Ponnet, K. The nitrate content of fresh and cooked vegetables and their health-related risks. PLoS ONE 2020, 15, e0227551. [Google Scholar] [CrossRef]
  59. Spokas, K.A.; Cantrell, K.B.; Novak, J.M.; Archer, D.W.; Ippolito, J.A.; Collins, H.P.; Boateng, A.A.; Lima, I.M.; Lamb, M.C.; McAloon, A.J.; et al. Biochar: A Synthesis of Its Agronomic Impact beyond Carbon Sequestration. J. Environ. Qual. 2012, 41, 973–989. [Google Scholar] [CrossRef]
  60. Ding, Y.; Liu, Y.-X.; Wu, W.-X.; Shi, D.-Z.; Yang, M.; Zhong, Z.-K. Evaluation of Biochar Effects on Nitrogen Retention and Leaching in Multi-Layered Soil Columns. Water Air Soil Pollut. 2010, 213, 47–55. [Google Scholar] [CrossRef]
Pollutants 06 00016 g001
Figure 1. Schematic diagram of potato plants grown at Kentucky State University HR Benson Research Farm in specially designed erosion plots used for collecting and measuring runoff and seepage water volume from soil amended with biochar or animal manure under various soil management practices.
Figure 1. Schematic diagram of potato plants grown at Kentucky State University HR Benson Research Farm in specially designed erosion plots used for collecting and measuring runoff and seepage water volume from soil amended with biochar or animal manure under various soil management practices.
Pollutants 06 00016 g001
Pollutants 06 00016 g002
Figure 2. Volume ± std error of runoff water (A) and volume of seepage water (B) from a potato field amended with six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow)—following a natural rainfall event. Statistical analysis was carried out among the soil treatments using analysis of variance (ANOVA). Bars labeled with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Figure 2. Volume ± std error of runoff water (A) and volume of seepage water (B) from a potato field amended with six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow)—following a natural rainfall event. Statistical analysis was carried out among the soil treatments using analysis of variance (ANOVA). Bars labeled with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Pollutants 06 00016 g002
Pollutants 06 00016 g003
Figure 3. Concentrations ± std error of Pb (A), Cd (B), and Ni (C) runoff water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 3. Concentrations ± std error of Pb (A), Cd (B), and Ni (C) runoff water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g003
Pollutants 06 00016 g004
Figure 4. Concentrations ± std error of Mn (A), Cr (B), and Mg (C) in runoff water collected from potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 4. Concentrations ± std error of Mn (A), Cr (B), and Mg (C) in runoff water collected from potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g004
Pollutants 06 00016 g005
Figure 5. Concentrations ± std, error of Cu (A), K (B), and Zn (C) in runoff water collected from potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Values labeled with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Figure 5. Concentrations ± std, error of Cu (A), K (B), and Zn (C) in runoff water collected from potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Values labeled with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Pollutants 06 00016 g005
Pollutants 06 00016 g006
Figure 6. Concentrations ± std, error of Pb (A), Cd (B), and Ni (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 6. Concentrations ± std, error of Pb (A), Cd (B), and Ni (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g006
Pollutants 06 00016 g007
Figure 7. Concentrations ± std, error of Cr (A), Cu (B), and Mg (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 7. Concentrations ± std, error of Cr (A), Cu (B), and Mg (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by a different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g007
Pollutants 06 00016 g008
Figure 8. Concentrations ± std, error of Zn (A), K (B), and Mn (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 8. Concentrations ± std, error of Zn (A), K (B), and Mn (C) in seepage water collected from a potato field under six soil management practices: biochar, chicken manure (CM), vermicompost (Vermi), sewage sludge (SS), non-amended soil (Control), and cow manure (Cow). Statistical analysis was carried out among the soil treatments for each metal using analysis of variance (ANOVA). Bars accompanied by different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g008
Pollutants 06 00016 g009
Figure 9. Heavy metal, ammonia (NH3), and nitrate (NO3) ions in runoff sediment and soil solution showing adsorption on clay (red brown dots), organic matter, and hydrogen ion replacement sites.
Figure 9. Heavy metal, ammonia (NH3), and nitrate (NO3) ions in runoff sediment and soil solution showing adsorption on clay (red brown dots), organic matter, and hydrogen ion replacement sites.
Pollutants 06 00016 g009
Pollutants 06 00016 g010
Figure 10. Concentrations of ammonia and nitrate in runoff water (A) and seepage water (B) collected down the field slope of specially designed field plots of white potato plants grown under six soil management practices: vermicompost (Vermi), sewage sludge (SS), biochar (Bio), cow manure (Cow), chicken manure (CM), and non-amended soil (Control) at Kentucky State University Research Farm. Bars accompanied by different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Figure 10. Concentrations of ammonia and nitrate in runoff water (A) and seepage water (B) collected down the field slope of specially designed field plots of white potato plants grown under six soil management practices: vermicompost (Vermi), sewage sludge (SS), biochar (Bio), cow manure (Cow), chicken manure (CM), and non-amended soil (Control) at Kentucky State University Research Farm. Bars accompanied by different letter(s) are significantly different (p ≤ 0.05) according to Duncan’s multiple tests for mean comparisons.
Pollutants 06 00016 g010
Pollutants 06 00016 g011
Figure 11. Nitrate concentrations in potato tubers (Solanum tuberosum var. Kennebec) under different soil amendments: biochar (Bio), chicken manure (CM), cow manure (Cow), sewage sludge (SS), vermicompost (Verm), and non-amended control (Control). Values are expressed as mg NO3− kg−1 fresh weight. Bars with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Figure 11. Nitrate concentrations in potato tubers (Solanum tuberosum var. Kennebec) under different soil amendments: biochar (Bio), chicken manure (CM), cow manure (Cow), sewage sludge (SS), vermicompost (Verm), and non-amended control (Control). Values are expressed as mg NO3− kg−1 fresh weight. Bars with different letters are significantly different (p ≤ 0.05) according to Duncan’s multiple range test.
Pollutants 06 00016 g011
Table 1. Animal manure and biochar applied under field conditions for growing potatoes, Solanum tuberssum var. Kennebec, their rate of application (A) and composition of animal manures and biochar (B) applied at KSU HR Benson Research and Demonstration Farm (Franklin County, KY, USA).
Table 1. Animal manure and biochar applied under field conditions for growing potatoes, Solanum tuberssum var. Kennebec, their rate of application (A) and composition of animal manures and biochar (B) applied at KSU HR Benson Research and Demonstration Farm (Franklin County, KY, USA).
(A)
Soil AmendmentsLbs. Plot−1Lbs. Acre−1Kg Hectare−1
Sewage sludge (SS)40.212000.002241.70
Chicken manure (CM)182.829090.9010,189.55
Cow manure (Cow)402.2120,000.0022,417.02
10% Biochar279.3613,888.8815,567.36
Vermicompost (Vermi)201.1110,000.0011,208.51
(B)
Soil
Properties		Sewage SludgeVermicompostCow
Manure		Chicken ManureBiocharNative Soil
(Control)
P (%)0.30 b1.23 a0.72 ab0.79 ab0.30 b0.14 c
N (%)0.47 b1.40 b1.75 b4.12 a0.56 c0.14 c
K (%)0.24 b0.54 ab1.22 a0.50 ab0.33 b0.26
OM (%)3.20 b7.60 a5.63 a6.19 a7.60 a2.45 b
C (%)3.6 c12.1 b26.1 a17.7 ab3.7 c1.5 c
C/N ratio7.6 c8.64 bc14.91 a4.29 c17.61 a10.71 b
P (mg kg−1)10.54 a0.23 c0.30 c2.56 b1.16 bc18.59 a
Pb (mg kg−1)19.49 b18.48 b22.95 a25.06 a19.92 18.97 b
Cd (mg kg−1)6.63 a6.65 a6.73 a6.82 a6.75 a0.66 b
Ni (mg kg−1)9.22 a0.34 c4.06 b6.27 b5.12 b6.83 b
Mn (mg kg−1)497.86 b213.46 c308.62 b201.2 c327.98 b1262.5 a
Cr (mg kg−1)19.43 a2.49 b3.31 b5.06 b3.2 b14.15 a
Mg (mg kg−1)2045.25 b1564.7 c2400.28 b6705.89 a2198.37 b847.71 d
Cu (mg kg−1)182.29 a1.05 d18.647 c37.793 b9.08 c8.58 c
K (mg kg−1)928.75 d506.2 d8279.41 b17,741.03 a2251.29 c292.55 d
Zn (mg kg−1)441.85 a122.1 ab58.73 b225.09 ab28.33 bc19.72 c
pH5.72 d6.57 c8.09 a7.33 b7.85 a6.29 c
Conductivity,
µS cm−1		3.79 e21.03 c2.74 e5.18 d410.33 a101.53 b
NO3–N (mg kg−1)567.35 b366.23 c84.45 d745.35 a92.63 d6.22 e
NH3+–N (mg kg−1)171.09 a93.5 c0.65 e144.04 b5.06 d1.05 e
Note: Soil amendments applied and mixed with native soil prior to white potato planting at the recommended rates of application [24]. Soil organic matter (SOM) was determined as the dry weight minus the ash content. The pH was determined using a glass electrode in soil/distilled water slurry (1:5 w/v). Elemental analysis was carried out using an inductively coupled plasma–optical emission spectrometer ICP/OES. Each value in the table is the average of three replicates ± standard error. Statistical comparisons between soil treatments were made using analysis of variance (ANOVA). Values accompanied by different letter(s) in each row are significantly different (p ≤ 0.05).
Table 2. Concentrations of metals expressed as µg mL−1 in runoff water (A) collected down the field slope using tipping bucket apparatus, seepage water (B) collected from the vadose zone, and runoff sediment (C) (Kentucky State University HR Benson Research Farm (Franklin County, KY, USA).
Table 2. Concentrations of metals expressed as µg mL−1 in runoff water (A) collected down the field slope using tipping bucket apparatus, seepage water (B) collected from the vadose zone, and runoff sediment (C) (Kentucky State University HR Benson Research Farm (Franklin County, KY, USA).
Soil Treatment(A) Heavy Metals Detected in Runoff Water, µg mL−1
MnZnPbCdNiCrCuMgK
Biochar2.1040.20.02830.030.030.030.0250.66332.638
Chicken Manure3.1960.0550.02670.030.030.030.02170.73833.128
Cow Manure2.730.0780.02830.030.030.030.02830.35830.703
Control2.3330.1680.02330.0250.0250.0250.01830.391.097
Sewage Sludge2.8350.0380.02670.030.030.030.02670.48673.3
Vermicompost1.4240.190.02830.030.030.030.01830.92674.327
Allowable Limits in Drinking Water0.05 [46]5 [46]0.00 [47]0.005 [47]0.10 [47]0.10 [47]1.3 [47] 40 [48]12 [49]
Soil Treatment(B) Heavy Metals Detected in Seepage Water, µg mL−1
MnZnPbCdNiCrCuMgK
Biochar0.0180.00060.00030.00050.000350.00030.00040.2720.0677
Chicken Manure0.0180.00150.00030.00050.000310.00040.00050.12440.097
Cow Manure0.0180.00030.00040.00050.000380.00030.00030.23130.0559
Control0.0110.00130.00040.00050.000330.00030.00040.18740.071
Sewage Sludge0.0120.00030.00040.00050.000310.00040.00020.16120.0585
Vermicompost0.020.00330.00040.00050.000360.00030.00070.23770.211
Allowable Limits in Drinking Water0.05 [46]5 [46]0.00 [47]0.005 [47]0.10 [47]0.10 [47]1.3 [47] 40 [48]12 [49]
Soil Treatment(C) Heavy Metals Detected in Runoff Sediments, µg mL−1
MnZnPbCdNiCrCuMgK
Biochar20.29874.310.5921.4651.1910.7262.0867.5491.81
Chicken Manure24.632106.71.0892.4611.4991.3674.16595.38154.65
Cow Manure8.25140.131.0581.6381.3610.9581.59924.7964.81
Control6.66493.252.0432.8482.2981.7232.28478.1108
Sewage Sludge24.57260.260.5191.2560.9910.5912.14773.6695.83
Vermicompost19.66761.981.3351.9831.2171.0421.62439.53104.39
Disclaimer/Publisher’s Note: The statements, opinions and data contained in all publications are solely those of the individual author(s) and contributor(s) and not of MDPI and/or the editor(s). MDPI and/or the editor(s) disclaim responsibility for any injury to people or property resulting from any ideas, methods, instructions or products referred to in the content.

Share and Cite

MDPI and ACS Style

Antonious, G.F.; Neupane, B.; Bordoh, E.K.; Dawood, M.H. Heavy Metal and Nitrate Mobility in Runoff and Seepage Water from a Field Amended with Biochar and Animal Manure. Pollutants 2026, 6, 16. https://doi.org/10.3390/pollutants6010016

AMA Style

Antonious GF, Neupane B, Bordoh EK, Dawood MH. Heavy Metal and Nitrate Mobility in Runoff and Seepage Water from a Field Amended with Biochar and Animal Manure. Pollutants. 2026; 6(1):16. https://doi.org/10.3390/pollutants6010016

Chicago/Turabian Style

Antonious, George F., Basanta Neupane, Edward K. Bordoh, and Mohammad H. Dawood. 2026. "Heavy Metal and Nitrate Mobility in Runoff and Seepage Water from a Field Amended with Biochar and Animal Manure" Pollutants 6, no. 1: 16. https://doi.org/10.3390/pollutants6010016

APA Style

Antonious, G. F., Neupane, B., Bordoh, E. K., & Dawood, M. H. (2026). Heavy Metal and Nitrate Mobility in Runoff and Seepage Water from a Field Amended with Biochar and Animal Manure. Pollutants, 6(1), 16. https://doi.org/10.3390/pollutants6010016

Article Metrics

Back to TopTop