The ground state electronic configuration for Sb is [Kr] 4d¹⁰5s²p³ and the covalent radius is 0.141 nm [2] . Filella et al. [3] noted that the primary Sb valence states are Sb⁺⁵, Sb⁺³, and Sb⁻³, with Sb⁵⁺ as the more commonly occurring oxidation state. Stability constants and other selected thermodynamic data are limited. Antimony compounds are considered priority pollutants in the United States and European Union and elevated antimony soil and sediment concentrations are primarily related to anthropogenic sources or high arsenic bearing sulfide ores. Stable anionic thiocomplexes include SbS₂²⁻ and SbS₄³⁻ [3] . Antimony has a wide range of uses including the manufacture of semiconductors, diodes, flameproof retardants, lead hardeners, batteries, small arms, tracer bullets, automobile brake linings, and pigments [3] .

In a detailed survey of the literature, Kabata-Pendias [4] reported the crustal average Sb concentrations are approximately 0.2 mg kg⁻¹, with argillaceous sediments showing slightly larger average concentrations of up to 4 mg kg⁻¹. Hou et al. [5] reported that the average Sb concentration ranges in Japanese soils were 0.83 ± 0.32 mg kg⁻¹. Stibnite (Sb₂S₃) is the most important antimony ore, followed by valentinite (Sb₂O₃). Other sulfide minerals include pyrargyrite, zinkenite, jamesonite, and boulangerite.

Filella et al. [3] reported an ordering for antimony species toxicity, which is antimonites (Sb³⁺) > antimonates (Sb⁵⁺) > organoantimonials. Kolesnikov et al. [6] evaluated the ecotoxicity of 23 metals, metalloids, and nonmetals in a Haplic Chernozem and proposed three hazard classes with Sb in class II (intermediate hazard class). Antimony was determined to be bound to relatively immobile Fe and Al oxihydroxides and to a lesser degree as antimony-organic substances [7] .

Antimony (III) and (V) ions undergo hydrolysis readily in aqueous solutions. In oxic water and soil conditions, the Sb(V) species are the major species [8] . Antimonic acid has been represented as H[Sb(OH)₆], Sb(OH)₅, or HSbO₃ and frequently forms polymers as the pH increases. Antimonous acid [Sb(III)] is frequently represented as SbO⁺, Sb(OH)₂⁺ in acidic media and as Sb(OH)₄⁻ or hydrated SbO₂⁻ in basic media. Antimony(III) chloride will form successive chlorocomplexes [8] . Under reducing conditions and in the presence of sulfur, the mineral stibnite (Sb₂S₃) will crystalize in acid soils, whereas in alkaline pH levels the mineral SbS₂ replaces stibnite.

Sb(III) prefers sulfur as a ligand; however, Sb³⁺ will form stable complexes with ligands such as citric, lactic, mandelic, and tartaric acids Antimony(V) forms complexes with polyhydric alcohols, polyhydric phenols, and citric, malic, and lactic acids. Antimonic acid (Sb₂O₅.nH₂O or H₃O₄Sb) forms complexes with low molecular weight acids that have oxygen-containing functional groups, including humic compounds. Pilarski et al. [9] investigated the adsorption of Sb(III) and Sb(V) with humic acid, demonstrating adsorption was effectively described using Langmuir isotherms. Bagherifam et al. [10] investigated high organic matter soils incubated with antimony and documented that humic substances were responsible for 63% of the extractable Sb. In the humic acid fraction, antimony was largely associated with the low molecular weight fulvic acids and correlated with total organic carbon and nitrogen contents [9] .

Thanabalasingam et al. [11] investigated antimony adsorption onto hydrous oxides of Mn, Fe, and Al. Capacity values decreased along the sequence MnOOH > Al(OH)₃ > FeOOH and adsorption on each substrate decreased gradually at pH values greater than pH 6. Factors influencing Sb sorption included substrate surface charge, chemical form of Sb and surface interactions. In oxic soil conditions antimony is present as Sb(OH)₆ and in anoxic soil conditions Sb(OH).

Tang et al. [12] and Vidya et al. [13] noted that mine waste leaching, the weathering of sulfide ores and shooting ranges using Sb adulterated lead-based ammunition are major Sb pathways for impacting soils and aquatic environments. Tang et al. [12] also documented that Sb toxicity in plants reduces i) root and shoot growth, ii) seed germination, and iii) yield potential. Furthermore, antimony induces chlorosis, reduces photosynthetic efficiency, membrane stability, and nutrient uptake, and increases reactive oxygen species. Vidya et al. [13] reported that the majority of plant uptake of Sb is confined to root tissues; however, some of the metalloid is translocated to the shoot. Inhibition of photosynthesis, modified root and leaf patterns, activation of antioxidant systems, and plant membrane disruption are some of the deleterious effects of antimony on plant growth and development. Bowen [14] documented that terrestrial plant uptake of antimony may result in a plant tissue concentration of approximately 60 µg kg⁻¹.

The ground state electronic configuration Bi is [Xe] 4f¹⁴5d¹⁰6s²p³ [2] . The covalent radius for Bi is 0.152 nm [2] . In a detailed survey of literature, Kabata-Pendias [4] noted the crustal average Bi concentrations are approximately 0.2 mg kg⁻¹, with argillaceous sediments showing slightly larger concentrations up to 4 mg kg⁻¹. In the United States, Govindaraju [15] reported that soil Bi averages range from 0.03 to 0.69 mg kg⁻¹. Hou et al. [5] reported average Bi concentration ranges in Japanese soils were 0.32 ± 0.12 mg kg⁻¹. Hou et al. [5] reported that average Bi distributions among the more dominant chemical fractions were noncrystalline (26%), peroxide extractable organic (26%) and metal organic (19%), residual (17%) and crystalline iron (12%). In Japan, Manaka [16] documented that Sb, and Bi were positively correlated with the amorphous Fe₂O₃.

Murata [17] investigated Bi solubility as influenced by pH and the presence of EDTA, citric acid, tartaric acid, L-cysteine, soil humic acids, and dissolved organic matter. Solution pH and the presence of citric acid, tartaric acid, L-cysteine, and soil humic acids influenced bismuth solubility. Kleja et al. [18] investigated the binding of Bi3+ to organic soil materials and noted that Bi3+ formed organic soil complexes, forming a dimeric Bi3+ complex.

The objectives of this investigation are: i) to estimate the soil abundances of antimony and bismuth across southeastern Missouri, and ii) to determine their potential soil profile distribution because of eluviation-illuviation and iron oxide abundances.

The study area in Missouri is located between the Mississippi River and the St. Francois River. The northern section consists of thin to thick loess mantles overlying primarily Precambrian igneous and Ordovician carbonate rocks. The southern portion is in the Mississippi River embayment and consists of floodplains and terraces that have coarse to fine sediments.

The climate is continental humid. The average daily January temperatures are 2 to 4˚C (35 to 39˚F), whereas the average summer temperatures are 25 to 26˚C (77 to 79˚F). The rainfall is reasonably well distributed, with the total annual precipitation averaging 1.20 m. The remnants of tropical storms from the Gulf of Mexico provide periodic intense rainfall events [19] [20] [21] .

Soils were selected from the following soil orders: i) Mollisols, ii) Alfisols, iii) Ultisols, iv) Entisols, v) Inceptisols, and vi) Vertisols. In total, 27 soil series were selected, many with multiple pedons. The soils used in this investigation were routinely characterized: i) to verify that the pedon was a member of the soil series, and ii) to provide routine soil chemical characterization. Laboratory analysis was performed only on the fine earth fraction, that is material finer than 2 mm. Standard routine methods included pH in water, exchangeable cations, total neutralizable acidity, and organic matter content by loss on ignition. These methods were performed by the soil testing laboratory at the University Missouri-Columbia Fisher Delta Center (Portageville, MO). Soil taxonomic classifications were from the United States Department of Agriculture official soil series descriptions [22] .

An aqua regia digestion was employed to obtain a near total estimation of elemental abundance associated with all but the most recalcitrant soil chemical environments. Homogenized samples (0.75 g) were equilibrated with 0.01 liter of aqua-regia (3 mole nitric acid: 1 mole hydrochloric acid) in a 35˚C incubator for 24 hours. Samples were shaken, centrifuged, and filtered (0.45 µm), with a known aliquot volume analyzed using inductively coupled plasma emission – mass spectrometry. Selected samples were duplicated and known reference materials were employed to guarantee analytical accuracy. A water extraction was performed to recover only the most labile or potentially labile fractions. A hot water extraction involved equilibrating 0.5 g samples in 0.02 L distilled-deionized water at 80˚C for one hour followed by 0.45 µm filtering and elemental determination using inductively coupled plasma emission – mass spectrometry. For the water extraction, selected samples were duplicated, and reference materials were employed to guarantee analytical precession. Simple statistics included mean, standard deviation (STD), coefficient of variation, and linear regression analysis were each performed using Excel.

The antimony concentrations ranged from very low concentrations to 0.65 mg kg⁻¹, whereas the bismuth soil profile concentrations ranged from very low concentrations to 0.92 mg kg⁻¹. For the Alfisols and Ultisols, the antimony soil profile distributions show greater concentrations in the illuvial horizons than the eluvial horizons, whereas for the antimony soil profile distributions for the majority of the Entisols, Inceptisols, Mollisols and the solitary Vertisol show rather uniform concentrations. Zhao et al. [23] noted that adsorption was an important phenomenon influencing antimony accumulation, with clay and iron oxyhydroxides important soil substrates.

The magnitude of the antimony and bismuth concentrations are within the concentration ranges documented for pristine, non-impacted (geogenic) soils [4] [23] [24] [25] [26] . In their review, Bolan et al. [24] did report that background and average antimony concentration values were 0.3 to 8.6 mg kg⁻¹. Thus, the sampled soils across southeastern Missouri are not inferred to be impacted by mining or other anthropogenic activities.

Tang et al. [26] recently described sources of soil antimony pollution and its environmental impact, and proposed remediation techniques, noting use of ferrous sulfate, phosphate amendments, clay minerals and biochar for adsorption, selected bacterial communities, and phytoremediation. Antimony pollution sources include mining, pharmaceutical manufacturing, vehicle emissions, plastic waste leaching, shooting ranges, and others [23] [25] [26] . The maximum concentration of antimony for drinking water is 6 ppb by the United States Environmental Protection Agency and 20 ppb by World Health Organization [26] . Vidya et al. [13] reviewed the influence of antimony on plant growth and development, including i) stunted growth, ii) reduced photosynthesis and accumulation biomass, iii) generation of reactive oxygen species, and lipid peroxidation. However, the literature addressing antimony and bismuth as environmental risks is a relatively recent activity.

Bolan et al. [24] noted that no definitive review of the biogeochemistry of antimony has completely described antimony mobilization, bioavailability, toxicity, and threats to environmental and human health. Bolin et al. [24] significantly reviewed recent literature of the biogeochemical processes influencing soil antimony, noting that clay minerals and the oxyhydroxides of Mn, Al and Fe were implemental in regulating antimony soil chemistry.