Metal and metalloid monitoring in water by passive sampling – A review

1 Introduction

Metals are ubiquitous elements that occur in all environmental compartments by natural or anthropogenic pathways. Over the last century, industrialization has proliferated, increasing the demand for the earth’s natural resources, which exacerbated the global problem of environmental pollution with organic and inorganic pollutants, radionuclides, or nanoparticles [1,2]. In particular, water contamination by the presence of contaminants harmful to human health is a critical problem all over the world. As a result of their persistence, bioaccumulation, biomagnification behavior, and potential for toxicity, trace metals and metalloids were among the water contaminants that underwent extensive studies [3,4,5,6].

Three-quarters of the elements in the periodic table are metals, and among them, only several elements are essential for living organisms. Non-essential elements (Cd, Pb, Hg, As, and Al), with no known biological function, proved to represent a risk to environmental and human health, even at relatively low concentrations [7,8,9], while some essential trace elements (Cu, Cr, Zn, Fe, V, Mn, Co, Ni, Se, Mo, etc.) became toxic in increased concentrations [10,11,12,13]. The tendency of metalloids to form covalent bonds with organic groups makes them exhibit toxicological properties [1]. Therefore, consistent chemical assessment of the presence of metals and metalloids in aquatic ecosystems has become a key topic, particularly in the context of the increasingly restrictive water quality standards [14]. This aspect increased interest in discovering easy-to-use and affordable alternatives to grab sampling. One of the most suitable substitutes to improve water quality monitoring and assessment is the use of passive sampling. Passive samplers (PSs) tools were intensively developed in the last two decades for a wide range of contaminants in waters, both for organic contaminants [15,16,17,18] and for inorganic contaminants, including metals and metalloids, which could be passively sampled [19].

Generally, PSs have a receiving phase that has affinity toward the compounds of interest [20]. Passive sampling rates are used to determine the concentration of the analyte in the test medium by means of the quantity of analyte accumulated in the receiving phase [21].

Since the working principles of PSs for metals and metalloids were previously assessed in several reviews, mainly focused on specific techniques such as diffusive gradient in thin film (DGT) [18,22,23,24,25], or polymer inclusion membranes (PIMs) [26,27], the main objective of this review is to provide a comprehensive update on the latest progress in passive sampling techniques for metal and metalloid monitoring, speciation, and fractionation in aquatic systems, and thus, it complements the other reviews existing in the literature. Some possible developments and future applications for water quality monitoring and assessment are also discussed in this article. Selected research articles published mainly within the past 4–5 years were examined in this review article, with a few exceptions that were not excluded due to their importance for the field.

2 Fundamental aspects of metal and metalloid passive sampling from water

The passive sampling tools are deployed in situ or in the laboratory for a specific period, to accumulate the analytes [28]. The transfer of contaminants from the aquatic environment into the sampler is based on the diffusion process determined by the difference in the chemical activity of the analyte between the tested medium and the sampler sorption phase [29]. No active transport or pumping is used for these samplers. The analytes pass to the sorption phase through a diffusion-limiting membrane, which allows the selective transfer of the analyte to enter the receiving phase [16]. The PS’s function can be in kinetic or equilibrium modes, depending on the exposure time, as presented in Figure S1 [30]. The analyte accumulates in the sorption phase until thermodynamic equilibrium is established between the sampler and the water. At the early period of PS deployment, analytes are absorbed at a rate directly proportional to their aqueous concentration; thus, the sampled amount of analyte is proportional to the time-weighted average of its concentration in the water phase during exposure. A first-order linear equation can be used to calculate the accumulation rate of analytes in the receiving phase, taking into consideration the deployment time [20]. Some factors, including temperature, matrix, and flow pattern of the aquatic system, can influence the transport of the analyte across the membrane, consequently affecting the accumulated amount of the target analyte [26]. PIM is a versatile technique having applications not only for analytical purposes in water monitoring, but also for industrial applications for wastewater decontamination in industrial processes to separate metal ions from the solution. Its functioning principles make this technique applicable for the removal of metal ions, dyes, and organic pollutants from aqueous solutions [27,29]. Heavy metals can be extracted from wastewater by different processes, such as membrane filtration, adsorption, and ion exchange. A PIM is composed of about 40% (w/w) base polymer, 40% (w/w) carrier, and 20% (w/w) plasticizer [30]. Since the polymer component of PIM ensures the mechanical strength of the membrane, the carrier oversees extracting the target heavy metals from the aqueous media and transporting these species inside the membrane structure. Thus, the carrier characteristics are critical, because they impact the process of the heavy metal’s transportation process. Depending on the carrier category, heavy metals are transported through the membrane in several ways [27]: (i) simple transport produced by metal solubility in the liquid membrane, which takes place until the equalization of the analyte concentrations in the feed and receiving phases; (ii) assisted transport that may arise by partitioning, complexation, or diffusion reaction. The metal ions from the feed phase are complexed by a carrier contained in the membrane and transported to the receiving phase. It takes place till the concentrations on both faces of the membranes are identical; (iii) counter-transport, which takes place versus the concentration gradient. It implies the transfer of metal ions to the receiving phase with the aid of a carrier, with the concurrent transfer of another substance (normally a hydrogen ion) in the inverse direction; (iv) coupled transport co-transport, which implies that a liquid substance is co-transported with an associated component in a process that ends when the concentrations between the receiving phase and the feeding phase are equal.

PIM is considered a greener replacement for the solvent extraction process, also being cost-effective, having higher selectivity, higher diffusion coefficient, and needing lower energy consumption if compared with the liquid membrane process [30,31,32]. Moreover, its preparation requires a minimum of hazardous substances, and in recent years, research has been conducted to include environmentally friendly biodegradable polymers into PIM structures [30]. The ion pair or the complex formed between the carrier and metal ions can diffuse across the membrane, and the analytes are then released to the receiving phase, and the carrier can repeat the transportation process. After the in situ exposure, the metal analytes are extracted from the sampler in the laboratory, and their concentration is then measured, often using a spectrometric technique.

2.1 Advantages and limitations of passive sampling

Passive sampling techniques offer several important advantages over discrete (grab) sampling methods. For example, grab sampling cannot indicate the variations of contaminants in time, while passive sampling can provide a time-averaged contamination during the exposure, being thus more suitable for estimating pollutant trends in water bodies [16]. Passive sampling permits the separation of the analyte from a large sample volume, sometimes with a complex matrix, in a small amount of a receiving phase [26]. Thus, the key advantage of this technique relies on its intrinsic specificity toward the analyte of interest. Moreover, only a fraction of the total analyte is sampled using this technique. In the case of metals and metalloids, only freely dissolved species and labile complexes can pass through the semipermeable membrane from the PS. Consequently, it is considered an improved proxy of the possible bioavailable concentration. Thus, in situ application reduces the potential contamination of the samples [31]. Also, passive sampling reduces the overall cost of the sample collection, simplifies the sample matrix, and improves the limits of quantification, thus offering the option to use less sensitive analytical methods for a varied range of pollutants [21,26,33]. PSs can combine the analyte collection, separation from the sample matrix, and preconcentration into a single step, simplifying the overall analytical procedure [26], being thus an inexpensive and unobtrusive alternative to more sophisticated analytical tools. Also, some PSs, such as PIMs, can be simply combined in compact and portable equipment to achieve online preconcentration and separation of the analytes [34]. Due to the relatively slow mass transfer across the separating membrane of a PS, this can be deployed in the studied environment over very long periods ranging from a few hours to a few weeks or months [26,35]. Thus, passive sampling offers an indication of average pollution levels over long periods of time. It requires no supervision, it is noiseless, and it can be used in hazardous environments. All the indicated advantages illustrate the “greenness” of passive sampling techniques for controlling emerging pollutants [36].

Among the limitations of passive sampling is the impact of the exposure environment on the uptake of substances by sampling tools. Therefore, the thickness of the diffusion layer between the receiving phase of the sampling tool and the examined aqueous media may change depending on the hydrodynamic settings [37,38,39]. Thus, the correct estimation of diffusion-layer thickness is fundamental for the accurate quantification of the analyte [40]. Another limitation of passive sampling that can be addressed is its integrative aspect: only the average value over the deployment time is measured, so the maximum or minimum concentrations during the deployment are not obtained. Also, passive sampling requires long-term sampling periods, usually in terms of days. Finally, the passive sampling of different target analytes requires diverse diffusion membranes or receiving phases [35]. Water chemical properties may also influence the analyte accumulation; thus, the validation for in situ application should be carefully carried out, in order to obtain reliable results [19].

When a resin gel is used as the receiving phase, an elution step is usually necessary to extract the analytes before instrumental determination. This is a crucial step requiring attention since it has been confirmed to be one of the most important contributors to method uncertainty [41]. For DGT with a Chelex binding layer, the elution factor of 0.8 is commonly used when the elution is performed in 1 mL of 1 M HNO₃. However, it is risky to use this value if the experimental conditions (volume of eluent, eluent type, concentration, elution time, type of analyte, type of binding layer, etc.) differ from normal conditions of use. Thus, the users should experiment with their own experimental elution factor to reduce the measurement uncertainty.

The experimental conditions, such as ionic strength medium, may also affect the diffusion through the limiting membrane from PSs [42]. Consequently, this aspect should also be taken into consideration, according to the experimental conditions (Scheme 1).

Scheme 1

Advantages and limitations of passive sampling.

Although the passive sampling technique has been recommended as an alternative to grab sampling within a legislative framework [35,43], future steps are still required to adopt this as an official sampling approach in routine environmental monitoring programs.

3 Types of PSs used for metal and metalloid measurement in water

A PS can function in equilibrium or kinetic modes. In the equilibrium mode, the target analyte reaches the equilibrium with the sampler receiving phase, whereas in integrative samplers, a kinetic sorption regime is created. A combination of the two sampling modes is also possible [35]. The common point of passive sampling is the interface between the sampled environment and the receiving phase. The transport of analyte through this interface may occur by diffusion through porous membranes or by permeation through nonporous membranes. The most used PSs for metal and metalloid measurement in water are presented in Scheme S1.

Even though the passive sampling technique requires a certain period to collect the target analyte for environmental monitoring, and usually two trips are needed for PS tool deployment and retrieval, it has an important advantage over grab sampling in which the sample is taken directly from the media at a specific point in time. The integrative characteristics of passive sampling reduce drastically the number of samples required for consistent monitoring. While passive sampling achieves the in situ accumulation of analytes during the deployment period in a single sample, with grab sampling, the collection of a much larger number of samples is required for similar results [35,44]. It permits economic monitoring, because of the reduced number of samples that are collected and subsequently prepared and analyzed in the laboratory. It implies that also much lower quantities of reagents are necessary for analytical determinations. Furthermore, due to the analyte concentration in the receiving phase, the volume of samples transported to the laboratory is largely reduced [44].

4 Passive sampling in water applied to chemical fractionation

Estimating the bioavailability of trace elements in aquatic systems is of major importance to efficiently protect aquatic environments. Consequently, chemical fractionation, speciation, and bioavailability assessment of metals and metalloids have received increased consideration in recent years. Following the international union of pure and applied chemistry recommendations [58], in this study, fractionation and speciation analysis are distinctly treated. Thus, even if in many studies dealing with passive sampling, the fractionation is referred as speciation, the articles in which labile species concentrations are separated from total concentrations based on diffusion according to size, solubility, bonding, or reactivity are considered to discuss about fractionation. In Scheme 2 is presented the use of PSs for the classification of analytes into chemical fractions.

Scheme 2

Fractionation analysis using PSs.

The diffusive gel of the DGT permits the measurement of metals in ionic form or metals weakly bound to small inorganic and organic complexes. It is generally called the “labile” fraction. Two different pore size diffusive gels were used to assess the bioavailable fractions of several metals [59]. Strong correlations between dissolved and DGT-labile concentrations were found for most metals. Caetano et al. [60] measured the labile chemical forms of Cd and Pb in waters by DGT-ICP-MS and anodic stripping voltammetry (ASV) and the total dissolved metal concentrations by inductively coupled plasma mass spectrometry (ICP-MS). The total dissolved concentrations of Ni were measured by ICP-MS and cathodic stripping voltammetry (CSV), while the labile forms of Ni were measured by DGT-ICP-MS. The study revealed that DGT could measure the potentially bioavailable concentrations of metal ions in aquatic systems and fit well for surveillance monitoring programs.

Traditional and passive sampling (DGT) was used to distinguish the dissolved and particulate metal and metalloid concentrations in estuary waters [61]. Fe, Pb, and Sn were predominantly found in the particulate matter, while the other studied elements were mostly in the dissolved fraction. For some metals, a significant correlation was found between the DGT-labile and dissolved fractions and metal bioaccumulation. The fractionation of REE and other trace elements (Ag, Tl, U, Y and Cs) in estuary water were also investigated by periodic grab sampling and DGT [62]. Most metals (Ag, Tl, U, Cs) were primarily found in the dissolved fraction, while REE and Y were found mainly in the particulate phase. A similar strategy, based on the assessment of total dissolved and particulate metal concentrations that traditional sampling combined with passive sampling by DGT, was used to investigate trace metals (Cd, Co, Cr, Cu, Ni, and Pb) fractionation and distribution in the Scheldt estuary [63]. This approach helped to determine that in the studied area, Cd and Ni were mainly in labile forms, while Cr and Pb were less bioavailable. DGT and biotic ligand model (BLM) were combined to investigate metal (Cu and Zn) bioavailability in water. While DGT-Cu did not correlate well with BLM Cu(ii), a positive correlation was found between DGT-Zn and BLM Zn(ii) [64]. Sans-Duñó et al. [47] used DGT devices containing different gels and/or resins to fractionate the Ni species water. It was found that the lability degree rises as the thickness of the diffusive gel increases. PIM was used to evaluate As(v) uptake in aquatic environments, and the results were compared with those for As(v) uptake by two filamentous fungi. The results were well correlated with the uptake of Aspergillus niger [65].

5 Passive sampling in water applied to chemical speciation

Only free metal and metalloid species are believed to be collected at the binding layer of PSs. This implies that metalloid complexes should dissociate to be accessible for uptake by PSs. Thus, PSs could separate species according to their mobility (diffusivity) and lability (the property to dissociate) [37]. Among the PSs, DGT was found to be the most widely used in the speciation of metals and metalloids probably due to its simplicity. In Scheme 3 is presented the schematic procedure for chemical speciation using passive sampling.

Scheme 3

Schematic procedure for chemical speciation using passive sampling.

Schmidt et al. [66] used DGT for the determination of organic and inorganic species of metals and metalloids in seawater. About 73% of Cd and 75% of Ni were DGT-labile, signifying that the two metals were mainly found in inorganic complexes. V was fully DGT-labile, confirming its inorganic speciation. A lesser amount of As was passively sampled by DGT. DGT also allowed the assessment of the labile species of REY in seawater, and an overall correlation among their bioavailability and DGT-measured metal concentrations [66].

Due to its significance for the environmental risk assessment, chemical speciation according to the oxidation state of elements is the most widely studied by passive sampling. Different oxidation states induce changed toxicity and mobility for elements such as Cr, As, Fe, and Hg [1]. DGT binding gels for the retention of oxyanion species (As, Se, Sb, V, Mo, Cr, P, W) containing ferrihydrite, titanium dioxide, and zirconium oxide were developed and used for the determination of oxyanions [67,68]. Even if these setups cannot usually directly differentiate the diverse oxidation states, the determination of labile oxyanion concentrations is significant because they can contribute to the speciation. By choosing an appropriate DGT configuration according to its binding of diffusion properties, a selective uptake of targeted species can be achieved. Viana et al. [69] measured assessed arsenic species (As(iii) and As(v)) at the soil/water interface using DGT tools with zirconium oxide and 3-mercaptopropyl (3-MP) were as binding phases to selectively fix As species. A Zn-ferrite (ZnFe₂O₄) binding gel in DGT was tested to assess inorganic and organic acids (monomethyl arsenic acid and dimethyl arsenic acid) in river waters and pore waters [70].

Mercury (Hg) is one of the intensively studied elements due to its health concerns. Its toxicity is strongly influenced by its chemical species. Methylmercury (MeHg) is known to have damaging effects on human and ecosystem health. Bratkič et al. [71] speciated labile inorganic Hg and MeHg in waters used as binding layer in the DGT 3-MP-functionalized silica resin. Total Hg was eluted from resin gels in 1 mL aqua regia and analyzed by sector field inductively coupled plasma mass spectrometer (SF-ICP-MS), while MeHg was extracted from the resin gels with KBr, H₂SO₄, and CuSO₄, followed by shaking, the addition of 10 mL of CH₂Cl₂, shaking again, and centrifugation. MeHg was analyzed by headspace coupled with gas chromatography and cold-vapor atomic fluorescence spectrometry.

Dissolved Hg and MeHg were also assessed using 3-mercaptopropyl DGT in river waters [72]. The concentrations measured by DGT were, in general, like those of Hg and MeHg in waters after grab sampling. However, the authors observed that DGT concentrations were overestimated in low-flow situations and highlighted that DGT measurements in surface waters are influenced by environmental conditions [72]. The usage of hyphenated analytical techniques for chemical speciation can be necessary to enable MeHg separation and inorganic Hg species after DGT sampling. Table 2 summarizes the newest research in chemical fractionation, chemical speciation, preconcentration, and sample matrix separation using passive sampling.

The recent literature on metal and metalloid fractionation and speciation is based mainly on the DGT technique. Many studies from recent years were carried out for on-site validation for this purpose. This demonstrates that the passive sampling is highly versatile and can be easily transferred for on-site chemical fractionation and speciation applications.

6 Passive sampling in water applied to time-averaged concentrations, preconcentration, and sample matrix separation

In passive sampling, the analyte is constantly accumulated within the receiving phase over the deployment time. During this period, analyte collection, preconcentration, and extraction from the sample matrix are combined into one stage. Thus, PSs play an important role in the pretreatment of environmental water samples. In PIM-type PSs, the receiving phase binds the target analyte after its transportation through a membrane. In this way, the analyte is extracted from the matrix and, if deployed for a sufficient period, it is preconcentrated. For example, Vera et al. [34] used a PIM to develop a flow analysis (FA) system for the matrix separation and measurement of arsenate in drinking water. A PIM device was used for the determination of Cu(ii) from aqueous medium [74].

Bretier et al. [75] used DGT as integrative passive sampling and high-frequency grab sampling to monitor dam flushing events on dissolved metal concentrations in river water. Also, DGT allowed the improved limit of quantification (LOQ) and the assessment of metals’ behavior during the flushing event. Coupling the DGT with ICP-MS permitted the measurement of Pb isotopes in water [76]. DGT provided a time-averaged concentration and preconcentrated Pb isotopes in situ, facilitating a precise measurement of 207Pb/206Pb and differentiation of various anthropogenic Pb sources in the environment. Guigues et al. [77] used a semipermeable membrane device to monitor metals and PAHs in river water. The LOQs improved by a factor 2–8, depending on the analyte.

7 On-site applications of passive sampling

The advantages offered by passive sampling, which include the measurement of the average concentration over time intrinsically, lower quantification limits, and provide cleaner sample matrixes, make these techniques extremely useful in many environmental studies [78]. In the last few years, PSs have been widely used for numerous metals and metalloids in the aquatic environment. Rougerie et al. [79] used DGT with three different receiving phases based on Chelex, TiO₂, and ZrO for Al accumulation from freshwater. They tested the linear accumulation capacities, which varied from ≈1 to ≈15 µg and which can be reached after a few tens of hours of field deployment if the water contains high Al concentrations. Thus, it is recommended to check the accumulation capacities to avoid the risk of DGT saturation. The testing showed that for Al accumulation, titanium dioxide DGT (DGT-Ti) should be used for pH >7, while Chelex-100 DGT and DGT-Ti should be used for pH ≤7.

Two DGTs with different resin gels were used for the determination of mercury and other elements in different environmental matrix [80]. It was found that the two DGT tools can bind metals in a wide domain of pH and ionic strength. Rodríguez et al. [81] studied the effect of seawater physicochemical parameters on the correlation between metal concentrations measured by DGT and those determined using discrete sampling in European seawaters. Furthermore, Pb- and Cd-labile fractions were determined by ASV, while CSV measured the total dissolved Ni concentration. Ten metals/metalloids (Al, Zn, Ni, Cd, Cu, Pb, Cr, As, Se, and Sb) in water were analyzed using grab and DGT passive sampling. Two binding phases were used: Chelex-100 for cations and zirconium oxide for As, Se, and Sb oxyanions. The DGT field deployments have generally good accuracies; however, for some elements, DGT sampling can offer only a trend of metal concentrations [82].

DGT and spot samplings were used to measure total dissolved metals (Cd, Ni, Pb, Cr, Cu, Zn, Al, Fe, Mn, and Co) and the in situ labile metal fraction in wastewater treatment plants [83]. DGT, biomonitoring with caged gammarids, and grab sampling of the dissolved phase were used and compared to monitor metal (Cd, Cr, Co, Cu, Mn, Ni, Pb, and Zn) contamination in water [84]. The contamination with Hg, As, Fe, Mg, Cd, P, and S of sediments from Lake Hongze was evaluated by in situ measurement using DGT [85]. N, P, Fe, and S in sediments from typical black odorous runoff in Wuxi and diffusion fluxes of contaminants across the sediment–water interface were assessed using the DGT technique [86]. Huang et al. [87] evaluated the suitability of the DGT technique for the measurement of REEs in the sediment porewater. Laboratory experiments showed an efficient binding of all REEs in Chelex-100 resin gel, with a linear uptake over 3–4 days. Chelex-100 DGT devices were subsequently deployed in estuarine and marine sediments for 3 days. The applicability of the DGT technique in monitoring wells was investigated [88]. Since this is a non-turbulent and low-flow water system, the influence of a diffusive boundary layer (DBL) was investigated. An average DBL thickness of 0.06 cm was reported. Bersuder et al. [89] developed procedures and guidelines for the sampling of transitional and coastal waters by synchronized collection of grab water samples and the deployment of PSs (DGT tools) for the measurement of trace element concentrations.

A PIM PS designed for the detection of Zn in freshwaters was also studied [90]. Radzyminska-Lenarcik and Witt [91] used PIM for the determination of Zn and Mn. A PIM sampler was used for the determination of Cd from seawater [92], Yaftian et al. [93] reported PIM usage for V determination, while Motsoane et al. [94] used PIM in the field for extraction of the target metals. Keskin et al. [95] developed PIMs with different structures for Pd recovery from solutions. A PS device enclosing hydrophobic and chelating phases, like a Chemcatcher^®, was used to measure metals (Cu and Mn) and PCBs and in water [96]. Urban roadside snow was collected and analyzed as a passive sampling means for metals and metalloids [97]. Table 3 presents several examples of current uses of PSs for on-site monitoring of metals and metalloids.

The research in the last few years has focused on the on-site application of passive sampling. Samplers that allow the simultaneous monitoring of as many contaminants as possible were preferred.

8 Perspectives

Passive sampling has been demonstrated to be an efficient and simple-to-use instrument for analyzing metals and metalloids in aquatic environments. Future research is expected to focus on improving the existing tools to bring this technique even more in compliance with the green and white analytical concept in analytical chemistry. Already, passive sampling is well in line with many of the 12 principles of Green Analytical Chemistry [98] since it allows to reduce the number of samples, it is performed in situ, it reduces the number of analytical operations, and it saves energy and quantities of reagents, reducing the amount of analytical waste. Additionally, toxic reagents are generally reduced, and thus, the operator’s safety increases. Moreover, many of the principles of White Analytical Chemistry [98,99,100] can be reached based on passive sampling: improves LOD and LOQ, greatly improves precision and accuracy if compared with grab sampling, reduces the consumption of energy and reagents, and generates waste. The in situ deployment requires minimal equipment and can be carried out by personnel without high qualifications.

Possible improvements in some devices might be addressed. Currently, many receiving phases or separation membranes are targeted for different analytes. The production of tools capable of enlarging the number of target analytes simultaneously determined can be a future development. The integration of PSs in analytical tools avoiding thus the elution steps should be a topic for future research. Considering that PSs decrease the LOQ and separate the analyte from the complex matrix, these can be coupled with miniaturized analytical instrumentation. In general, PSs are still expensive devices for routine analysis. The development of cost-efficient tools for passive sampling is mandatory to implement this technique in routine laboratories.

9 Conclusions

Numerous studies published within the past 4–5 years have shown that passive sampling is a powerful tool for monitoring metals and metalloids in aquatic environments. Due to the accumulation in time from the sampled environment into the sampler, they can improve the limits of quantification and separate metals and metalloids into a simpler matrix before instrumental determination. Their common point is the interface between the sampled environment and a receiving phase that is assured by a diffusive media. Recent developments were reported for PSs, such as DGT, PIMs, SLMDs, “octopus” samplers, and active-PSs. Many of the articles published in this period proposed the development of new binding phases or separation membranes, aiming to enlarge the number of analytes, including their fractionation and speciation. Chemical fractionation, speciation, and bioavailability assessment have received increased consideration in recent years since only free element species are supposed to be collected by PSs. Moreover, DGT binding gels for selective retention of oxyanion species (As, Se, Sb, V, Mo, Cr, P, W) were created for the determination of oxyanions. Even if they cannot directly discriminate the diverse oxidation states, they finally contribute to the speciation. Also, methods for Hg speciation as total Hg and MeHg in waters using the DGT were developed. Passive sampling is becoming a widely recognized practice, demonstrated by numerous articles dealing with the in situ applications of this approach. Step forwards were made to increase the robustness, accuracy, and precision of the method, and collaborative studies were carried out to validate in-field applications. This tool can significantly contribute to bringing the analytical techniques used for metals and metalloids in water samples to align with the analytical chemistry’s green and white analytical concept.