Trace Element Speciation in the Environment: A Mini-Review

Cristina M. Couto^1,2,3* and Agostinho A. Almeida³

1. Associate Laboratory i4HB - Institute for Health and Bioeconomy, University Institute of Health Sciences - CESPU, 4585-116 Gandra, Portugal.

2. UCIBIO - Applied Molecular Biosciences Unit, Forensics and Biomedical Sciences Research Laboratory, University Institute of Health Sciences (1H-TOXRUN, IUCS-CESPU), 4585-116 Gandra, Portugal.

3. LAQV/REQUIMTE, Department of Chemical Sciences, Faculty of Pharmacy, University of Porto, Porto, Portugal.

Received: 01 May 2024; Processed: 09 June 2024; Accepted: 15 June 2024

Citation: Couto Cristina M. and Almeida Agostinho A. “Trace Element Speciation in the Environment: A Mini-Review.” J Environ Toxicol Res (2024): 106. DOI: 10.59462/JETR.1.2.106

Copyright: © 2024 Cristina M. Couto. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

Keywords

Trace elements speciation • Chromium • Mercury • Arsenic • Selenium • Thallium

Introduction

According to current IUPAC definitions, “speciation” corresponds to the “distribution of an element amongst defined chemical species in a system and “speciation analysis” refers to “analytical activities of identifying and/or measuring the quantities of one or more individual chemical species in a sample” [1]. In the context of trace elements (TE) chemistry, “species” are the specific forms of a given element in terms of isotopic composition, electronic or oxidation state, complex or molecular structure and phase [2].The importance of characterizing and determining the different chemical species of a given element (“speciation analysis”), is widely recognized as being of fundamental importance in environmental chemistry. Chemical speciation may be implemented to include functionally defined speciation (e.g., determination of species that are available to plants or present in exchangeable forms), and/or operationally defined speciation, which refers to the determination of extractable forms of an element [3].

The main chemical forms in which TE can occur in the environment include free ions, different oxidation states, metallic nanoparticles and various complexes (organic and inorganic) resulting from the interaction with reactive groups present in the medium that bind and stabilize them [4]. This chemical speciation is affected by a wide range of physicochemical factors, such as Arsenic temperature, pH, redox potential, greater or lesser presence of reactive chemical species (e.g., organic and inorganic complexing ligands), colloidal matter and other surfaces with adsorption potential [4]. The partitioning and biogeochemical cycling of TE in the different environmental compartments from the loading source depend on the physicochemical properties of their original chemical species and the particular physicochemical of the receiving ecosystem [5]. Then, the toxicokinetics and toxicodynamics of TE also depend on their chemical form, which often greatly affects the organism’s ability to regulate and/or store them, therefore exerting a direct influence on the expression of their toxicity [5].

Thus, because TE speciation can have a huge impact on the transition kinetics, partitioning, deposition, and potential resuspension in environmental compartments and on their ecotoxicological effects on living organism, TE speciation analysis has been increasingly recognized as of crucial importance [4].The most relevant chemical forms to be studied in environmental samples are free ions, kinetically labile (mostly weakly bound) and non-labile (mostly strongly bound) complexes, and the colloidal fraction [6]. Several studies have demonstrated that free ions and some labile complexes constitute the bioaccessible fraction of an element, being its chemical species most easily assimilated by organisms and, in the case of heavy metals, their toxic fraction [7]. Chemical speciation analysis has therefore also become a fundamental tool for assessing the environmental and ecotoxicological risks posed by TE and is currently the core of most metal pollution studies [8].

The importance of speciation analysis is demonstrated by the growing number of publications focused on this type of study [8-11]. Speciation analysis has benefited greatly from the development of highly selective and sensitive analytical methods, which have been implemented mainly through the coupling of chromatographic techniques with sensitive element-specific detectors [12]. Non-chromatographic methods have also been developed, although with a more limited separation capacity, generally limited to specific chemical species [12]. In both approaches, it is important to continue the development of improved methods, which in particular must consider their environmental impact as a crucial parameter to be optimized, as their main purpose is precisely environment assessment [8]. Novel analytical approaches must focus on avoiding or reducing the undesirable environmental side effects of chemical analysis, while maintaining the required classical analytical features of selectivity, sensitivity/detectability, precision, and accuracy [8]. Several recent reviews on TE speciation methods are available in the literature [8,13,14].

In addition to the scarcity of accurate speciation analytical methods, the lack of certified reference materials (CRM) and widely accepted analytical protocols are major limitations in the field of TE speciation analysis in environmental contexts [15]. The lack of sample collection, conservation, and handling protocols that demonstrably ensure the preservation of the original chemical species between sample collection and analysis is also a much felt gap. This mini-review aims to summarize the latest developments in the field of TE speciation analysis in environmental matrices (water, soil, sediment, biota), with a focus on data from the last eight years and relating to chromium, mercury, arsenic, selenium and thallium. The first four elements have been extensively studied, while thallium is still a somewhat recent arrival to this list of problematic environmental TE. Their most relevant environmental occurring forms in plants, soils, fresh and sea waters, and new or modified analytical techniques for their characterization and determination are presented and discussed in this mini-review.

Discussion

Chromium

Chromium (Cr) has a complex electronic and valence shell structure, with great potential for conversion from one oxidation state to another, existing in several valence states, from 0 to VI [16]. Trivalent (Cr(III)) and hexavalent (Cr(VI)) chromium are the most stable and the predominant oxidation states in the environment [16,17]. Cr(III) occurs mainly as cations (Cr³⁺), while Cr(VI) generally occurs as oxyanions (chromate: CrO₄²⁻, hydrochromate: HCrO⁴⁻, and dichromate: Cr₂O₇²⁻). Cr(III) is relatively stable and nontoxic at circum-neutral pH (6.5–8.5) due to the formation of insoluble hydroxides and oxide compounds and strong complexation with environmental chemicals; it has a strong tendency to form kinetically inert hexacoordinate complexes with water, sulfate, ammonia, organic acids, halides, and urea [16]. Soluble Cr(III) species form complexes with low molecular mass (e.g. succinic, oxalic, malic, citric, malonic) and high molecular mass (e.g., fulvic and humic) organic acids [18]. The oxidation states Cr(II), Cr(IV) and Cr(V) are unstable forms produced during the redox processes of Cr(III) and Cr(VI); under anoxic conditions, Cr(III) is typically the predominant form; under oxic conditions, anionic Cr(VI) species tend to predominate [16].These are soluble and mobile across the full pH range, presenting high mobility in water systems, which makes Cr(VI) much more toxic than Cr(III) [16]. Microorganisms can reduce Cr(VI) [16].

In plants, Cr is a potentially toxic TE, without any essential metabolic functions [19,20]. There is no specific transporter for Cr uptake by plants and it enters plants mainly through specific and non-specific essential ion channels [19]. Cr accumulates mainly in plant root tissues, with very little translocation to the shoot [19]. Cr causes harmful effects on various biochemical, physiological and morphological processes. It exerts phytotoxicity by interfering with nutrient uptake and photosynthesis, disrupting plant growth [19]. In particular, it induces increased generation of reactive oxygen species (ROS), causing lipid peroxidation and disrupting antioxidant activities [21]. Most soils contain Cr levels in the range of 15-100 μg g^-1 [22]. In freshwater Cr ranges from 0.1 to 117 μg L^-1, while seawater contains Cr concentrations below 0.5 μgL^-1 [22,23]. The range of airborne chromium levels in Europe is 0–3 ng m^-3 in remote areas, 4–70 ng m^-3 in urban areas, and 5–200 ng m^-3 in industrial areas [22,24]. Cr concentrations in plants are generally low, typically ranging from 0.02 to 0.2 μg g^-1 (dry weight) [16]. Environmental Cr speciation studies have been reported in water, soils and sediments [25,26], soil-plant systems, [19] and biota [27].

There have been several reports of new or modified techniques for Cr speciation. Two reviews specifically address the pros and cons of the analytical speciation methods [13,16]. Selected works relating to the environmental monitoring of Cr are referred in Table 1. Electrodialytic separation of Cr(VI) and Cr(III) followed by determination of Cr by graphite furnace atomic absorption spectrometry (GFAAS) was applied by Nugraha et al. on monitoring a soil extraction process [28]. Cr(III) and Cr(VI) from spring, river, and sea waters were selectively analyzed using a dispersive micro-solid phase extraction (D-μ-SPE) procedure with magnetic graphene oxide as sorbent and flame atomic absorption spectrometry (FAAS) as detection technique [29]. The extraction procedure used strong acids and organic solvents, albeit in small volumes, minimizing its environmental impact. One paper describes a procedure based on magnetic solid phase extraction (MSPE) using magnetic nanoparticles (MNPs) functionalized with iminodiacetic acid combined with GFAAS [30]. The method was applied to determine Cr(III) and Cr(VI) in lake and river water samples. A recent paper by Saiz et al. [31] provided a complete snapshot of the Cr(VI)-Cr(V)-Cr(III) speciation.

Table 1. Trace element speciation analysis – selected recent works

The authors describe a combination of ultravioletvisible (UV-Vis) and electron paramagnetic resonance spectroscopy (EPR) techniques for Cr speciation after a coupled adsorption and reduction process in zirconium metal-organic frameworks (MOF). UV-Vis spectral fingerprints of Cr(VI) and Cr(III) showed distinctive features, which allowed qualitative information about the Cr coordination environment. EPR spectroscopy proved to be highly sensitive in detecting transient Cr(V), in addition to being highly effective in detecting Cr(III), either as isolated or clustered species. Huang et al. [27] reported a study on the location and speciation of Cr in a Chinese plant (Coptis chinensis) with the aim of better understand the mechanisms of Cr accumulation and transport and, ultimately, contributing to minimizing Cr transfer to the food chain. The authors used synchrotron radiation microscopic X-ray fluorescence (SR-μXRF) and laser ablation inductively coupled plasma mass spectrometry (LA-ICP-MS) to spatially localize Cr in the plant, X-ray absorption near-edge structure (XANES) spectroscopy for Cr speciation analysis and inductively coupled plasma mass spectrometry (ICP-MS) to determine subcellular Cr levels. XANES data showed that Cr(VI) could be reduced to Cr(III) when the plant was grown with Cr(VI), and provided a novel conclusion: that this plant contained elemental Cr. Another work aimed to develop an analytical method capable of simultaneously providing molecular and elemental information from a single sample injection, which was achieved by coupling high-performance liquid chromatography/diode-array detection (HPLC-DAD) with ICP-MS [32]. Cr(VI) and Cr(III) in water have also been differentiated based on their distinct surface-enhanced Raman scattering (SERS) spectral features [33].

Mercury

In the environment, mercury (Hg) exists in different forms, including metallic/elemental Hg (Hg0), inorganic Hg (HgS, HgCl₂), and organic Hg – mainly as methyl Hg (MeHg), but also ethylmercury (EtHg) and phenylmercury (PhHg) [34,35]. Due to its high volatility and susceptibility to oxidation, elemental Hg is the predominant species in the atmosphere, while Hg(II) is the predominant species in water, soil and sediments, and MeHg is the main species in biota [36,37]. The toxic effects of MeHg can be significant due to its bioaccumulation and biomagnification (up to 10⁶) in the aquatic food chain, high affinity for macromolecules and slow metabolism [38]. Several environmental parameters (pH, organic matter, temperature, light, flooding conditions, ionic activity, redox potential, dissolved oxygen, sulfide levels, suspended solids and microbial activity) influence Hg speciation and bioavailability [39]. Mobile Hg-humic or fulvic complexes tend to be susceptible to methylation, whereas Hg bound to larger organic matter (OM) particles hinders methylation. Specific OM compounds can promote Hg²⁺ methylation by increasing bacterial activity [40,41]. OM can also facilitate Hg²⁺ methylation by inhibiting HgS precipitation or enhance HgS dissolution, thereby providing available Hg²⁺ for methylating microorganisms [14].

Soil microorganisms can also transform inorganic Hg(II) into Hg0 by the action of Hg reductases, which are found in bacteria such as Pseudomonas sp., S. aureus, Thiobacillus ferrooxidants, Streptomyces sp. and Cryptococcus sp. that have the Mer gene [41,42]. Sulfate-reducing bacteria are the main agents responsible for MeHg production in coastal sediments [43]. OM can also influence MeHg production by reducing the amount of bioavailable Hg(II) in the dissolved phase and stimulating the activity of methylating bacteria, providing substrate for mineralization [44]. Plants efficiently absorb Hg through their roots, since Hg is highly soluble in water and easily converted into the gaseous phase. Most of Hg absorbed by plants is retained in the roots, and only a small amount is translocated to the aerial part [45]. Hg causes phytotoxicity and impairs numerous metabolic processes even at low amounts, resulting in ROS production, lipid membrane oxidation, DNA and protein damage, inhibition of photosynthesis and growth retardation [14,46].

Uncontaminated freshwaters generally contain total mercury (Hg(t)) levels <5 ng L^-1 and background soil concentration is typically between 0.003 and 4.6 μg g^-1 [47]. In uncontaminated sediments, Hg(t) levels are typically between 0.2 to 0.4 μg g^-1, however, in sediments close to industrial and urban areas, the Hg(t) concentration can reach 100 μg g^-1 and MeHg can reach 0.1 μg g^-1 [47]. In recent years, a large number of papers have been published on Hg speciation in various environmental matrices (Table 1). Yang et al. [48] used graphene oxide-bounded silica particles as solid phase extraction (SPE) adsorbent for the online preconcentration of Hg (II), MeHg, and EtHg in drinking water, followed by selective determination by HPLC-ICP-MS. In another work, zwitterion-functionalized polymer microspheres (ZPMs) were used as a core adsorbent in an online SPE procedure for the enrichment of mercury species (inorganic, MeHg and EtHg) in environmental waters (surface and marine), also followed by selective determination by HPLC-ICP-MS [49].

MSPE has also been used for Hg analysis. The magnetic Fe₃O₄@SiO₂@ glycidyl methacrylate (GMA)-S-SH nanoparticles were used for the extraction of Hg(II), MeHg and PhHg from water and soil samples, followed by HPLC-ICP-MS determination [50]. The developed method avoids the use of organic solvents and requires only a small volume of dilute acid, demonstrating great potential for routine environmental analysis. In a similar analytical procedure, magnetic Fe₃O₄@SiO₂@ γ-mercaptopropyltrimethoxysilane (γ-MPTS) nanoparticles were prepared and used for the speciation of mercury (Hg (II), MeHg and PhHg) in environmental water, wastewater, tap water, and fish samples [51].

Thallium

In recent decades, thallium (Tl) has become a “technologycritical element” due to its increasing technological use [52]. As Tl present in non-contaminated environmental compartments and that of anthropogenic origin typically exhibit different isotopic signatures, this has been used as a strategy to fingerprint and identify anthropogenic sources of Tl and understand its environmental processes [53]. The two stable isotopes are ²⁰³Tl (29.5%) and ²⁰⁵Tl (70.5%). Tl has two main oxidation states, Tl(I) and Tl(III), both of which are highly toxic to humans and animals, microorganisms and plants, and is more toxic to mammals than Cd, Pb, and even Hg [22]. Most lake and river waters have low levels of Tl, in the range of 5-10 ng L^-1. In seawater, Tl is predominantly as the free ion Tl⁺ and as subordinate dissolved complexes TlCl⁰, and ranges from 0.2 to 20 ng L^-1 [54]. Tl concentrations in continental and oceanic crust and rocks are generally less than 1 μg g^-1 [22]. Some soils have a naturally high background level, but in general Tl concentrations in uncontaminated surface soils range from 0.1 to 2 μg g^-1 [54,55]. Tl is a non-essential element for plants and its average concentration in land plants is typically < 0.1 μg g^-1; the average Tl content in air does not exceed 1 ng m^-3 [11, 54].

The toxicity of Tl-based compounds is due to the similarity between Tl⁺ and K⁺ cations, which results in the disruption of K⁺-associated metabolic processes [56]. Tl(I) is expected to be the dominant species in almost all environmental systems, in equilibrium with atmospheric oxygen and in the absence of complexing agents. From a thermodynamic point of view, the conversion of Tl(I) to Tl(III) is expected only in the presence of extremely strong oxidants and high alkalinity [52], but photochemical reactions in surface waters or microbiological processes may lead to the oxidation of Tl(I) to Tl(III) even under even moderate oxidizing conditions[57]. Recent studies have also highlighted the possible role of Fe(III) and As(V) in the formation of Tl(III) compounds in acid mining drainage [52]. Tl(III) is approximately 50,000 times more toxic than Tl(I) on a free-ion basis [58]. In the terrestrial environment, Tl is usually bound to the soil matrix, which considerably limits its transport, although dissolved Tl (soluble salts) is susceptible to leaching and can be introduced into the aquatic environment [55]. Tl sorption in soils is commonly attributed to ion exchange reactions on Mn oxides, clays, or OM. A very recent paper describes the use of recycled Al beverage can powder (AlCP) in Tl soil remediation. AlCP was used as a substitute for zero-valent Al to drive the Fenton-like reaction that induced the oxidation of Tl(I) and subsequent precipitation of Tl(III) via medium alkalinization [59].

The high concentration of Tl in shallow soil poses a threat due to possible uptake by plant roots and accumulation in plant biomass [55]. The extent of the role played by natural organic matter (NOM), such as humic and fulvic acids, in the bioaccessibility of Tl is still poorly known. One study investigated the complexation of Tl(I) by a purified humic substance, used as a proxy of NOM [60]. The experiments were carried out with the Donnan membrane technique to separate the free Tl(I) ion from its complex form in the bulk solution. Tl has been reported to be widely distributed in plant tissues. It is thought that most of the uptakenTl travels along nutrient pathways to the cell cytosol and storage vacuoles [61]. Because Tl⁺ and K⁺ have similar uptake pathways, they can both easily accumulate in plants from the soil, affecting antioxidant enzyme activity and photosynthesis. Tl⁺ can interfere with Na⁺/K⁺-ATPase and pyruvate kinase, inducing oxidative stress, which leads to damage to cell membranes, proteins, lipids and chloroplast pigments, ultimately leading to cell death [62].

Due to its generally very low concentrations, knowledge of the distribution, enrichment, and environmental risks of Tl in different environmental matrices has benefited greatly from the major developments in atomic spectrometry techniques in recent years, especially inductively coupled plasma mass spectrometry (ICP-MS), due to its very high sensitivity, ease of coupling to chromatographic systems for speciation analysis and isotopic analysis capacity [63-66]. Thus, as an example, it was possible to demonstrate that soils and sediments contaminated from industrial emissions exhibited clearly different Tl isotopic compositions compared to the natural background, allowing source tracking and also new insights into the biogeochemical cycle of Tl in soil [66]. Other relevant studies are referred to in Table 1. A fast, reliable and relatively simple procedure for the separation of Tl(I) from Tl(III) was described by Lopez-Garcia [67], using a dispersive liquid-liquid microextraction step to isolate the ion pair formed by Tl(III) and cetylpyridinium chloride in the presence of bromide. In this way, Tl(I) remained in the aqueous phase, which was subjected to analysis, and the concentration of Tl(III) was obtained by difference.

The procedure was applied to Tl speciation in tap, spring, river and sea waters. A magnetic metal-organic framework nanocomposite with MNPs was used for Tl speciation analysis in well, sea, and wastewater samples [68]. Tl(I) was selectively retained by the sorbent, and total Tl was determined using hydroxylamine hydrochloride to reduce Tl(III) to Tl(I). Another paper described the selective and highly sensitive determination of Tl(I) and Tl(III) (as well as total Tl) using a novel ligandless mixed micelle cloud point extraction and GFAAS [69]. A detection limit of 15 ng L^-1 was obtained.

Arsenic

Arsenic (As) rarely occurs in the free state and is typically found in combination with sulfur, oxygen, and iron, so more than 100 arsenic compounds are present in the environment and biological systems [10]. As occurs in four oxidation states, –3 (e.g. arsenic hydride or arsine gas), 0 (e.g. crystalline As), +3 (arsenite: AsO₃³⁻), and +5 (arsenate: AsO₄³⁻), and in a wide variety of organic and inorganic As compounds [70-72]. Trivalent As is 60 times more toxic than oxidized pentavalent arsenate, inorganic As compounds are about 100 times more toxic than the methylated forms [70]. Organic species are common in OM-rich waters and enzymatic methylation of inorganic As occurs through the action of microorganisms (including fungi and bacteria), algae, and phytoplankton, with the formation of organo-As compounds such as monomethylarsonic acid (MMA), dimethylarsinic acid (DMA), trimethylarsine and arsenobetaine [73].

This methylation process can occur in soil, sediments and water bodies, particularly under anaerobic conditions [73]. On the other hand, organo-As can generate inorganic As through hydrolysis and biological degradation [74]. As precipitated in soils, sediments and wetlands can be released back into the environment by bacteria, fungi, and other microorganisms. The ratio of As(III) to As(V) in groundwater is a function of the OM content, biological activity and oxygen levels [72]. In seawater, inorganic As is usually present in the form of As(III) and As(V) [75]. The uptake and accumulation of As by plants varies according to habitats and plant species [76]. Methylated As and As(III) move through the nodulin 26-like intrinsic proteins aquaporin channels, while As(V) is taken up via phosphate transporters [77]. Some plants (e.g. Pteridaceae family) exhibit hyperaccumulation behavior toward As in aboveground tissues [78]. Still, in general chelation leads to detoxification of As(III) through complexation with thiolrich peptides [79].

As is present mainly as arsenobetaine in fish and crustaceans. Arsenosugars are the predominant species in algae, with relevant presence also in bivalves and spo