Humic substances influence on the distribution of dissolved iron in seawater: A review of electrochemical methods and other techniques

2.1 Historical developments of Fe–Lt detection by electrochemical techniques

The pioneering analysis of trace metal concentration utilizing electrochemical techniques was conducted to determine zinc and cadmium in the Dead Sea through anion stripping voltammetry (ASV, Figure 2) [29]. This method was subsequently extended to other trace metals, including lead and copper (Cu) [33], and bismuth and antimony [34], pushing the detection limits far below the sensitivity of prevailing analytical procedures such as spectrophotometry, specific glass or membrane electrodes, and polarography [35]. However, due to mercury toxicity, certain regulations have constrained the advancement of mercury electrode–ASV for detecting trace metal concentrations. Consequently, since 1973, the evolution of ASV has bifurcated into two primary directions (Figure 2): one focusing on replacing mercury with alternative electrodes for trace metal detection (e.g., mercury-coated graphite electrode [34], rotating glassy carbon electrode mercury plated [36,37], mercury film electrode [38], and a twin gold disc electrode [39], and the other emphasizing the detection of low-level and organic speciation of trace metals in natural and seawaters [30,40,41,42]).

The utilization of electroanalytical methods for discerning the organic speciation of trace metals in natural water was initiated in 1974 [30]. This included the assessment of the Lt concentration, conditional stability constant (K′), and inorganic metal ions via ASV (Figure 2). Chau et al. employed differential pulse ASV (DPASV) to measure labile copper in lake water [30], while Nürnberg and Raspor used DPASV with nitrilotriacetate as a model ligand to examine dissolved organic matter (DOM) [41], estimating the Lt concentration and stability of its chelation with cadmium, zinc, and lead (Figure 2). However, ASV is limited in its ability to only identify the presence of complexation with metals and Lt and does not permit quantification of the concentrations of natural ligands and K′ in seawater.

The concentration of Fe was initially determined by adsorptive cathodic stripping voltammetry (AdCSV) starting from 1991 [43]. This method was refined for detecting processes [23,32,44] and was applied to study the distribution of Fe in various seas, including the western North Atlantic Ocean [45]. Over two decades, scientists have modified different reagents (competitive ligands and buffers) to enhance performance and adapt to various natural waters for the determination of DFe concentration [46,47,48], particularly for detecting levels lower than picomolar of DFe [49]. The use of AdCSV for detecting Fe in seawater has demonstrated significant advantages, including a very low detection limit and minimal pretreatment requirements.

The method of AdCSV with ligand competition, developed by Constant M. G. van den Berg post-1984, began to identify the complexation with metals and ligands, thereby addressing several associated problems (Figure 2) [50,51,52]. Subsequent to 1994 [5,32], the technique of competitive ligand equilibration (CLE)–AdCSV has been employed to ascertain the concentration of Fe-binding Lt and their K FeL ′ . Both these parameters are crucial in understanding their influence on Fe biogeochemistry [16]. CLE–AdCSV is an indirect approach that involves titrating natural ligands with Fe in the presence of a known competing artificial ligand to yield the thermodynamic characteristics of organic Fe binding ligands, Lt concentration, and K FeL ′ [53,54]. This method has enabled scientists to determine the distribution of trace metal complexation by organic ligands in both open oceans and coastal waters [8,47,55,56].

In an effort to elucidate the characteristics of DFe complexation and labile fractions, researchers have undertaken numerous initiatives across various domains. Initially, to classify complexation strength, DFe complexation in seawater is often divided into two primary binding site categories: strong binding sites (L₁-type, log K _Fe′L ≥ 12–13) and weak binding sites (L₂-type, 12 ≥ log K _Fe′L ≥ 10, and potentially <10) throughout the water column [5,16,25,57]. Some studies even proposed three or four binding site categories [58]. Weak binding ligands are frequently associated with HSs. Additionally, the computational methodology for CLE–AdCSV has undergone significant refinements, encompassing both the calculation procedure [59] and method intercomparisons from 15 laboratories [60]. Moreover, the assessment of DFe complexation employs a diverse range of analytical window analyses [54,61,62], all striving to closely align with the actual complexing conditions. In summary, despite challenges related to data inter-comparability arising from varied reagents (competitive ligands and buffers), instrumental configurations, and fitting models [60], a wealth of measurements has been amassed [63], revealing the basin-scale distribution of Lt across the world’s oceans.

Despite significant advancements in electrochemical methods for measuring trace metal speciation, the determination of Lt and HSs using these methods presents limitations due to their inherent constraints [64,65]. It has been found that CLE–AdCSV titrations measure the bulk properties of a heterogeneous ligand pool in seawaters, adding excess ligands rather than the actual ligands present [66]. Although there is some debate about the oversimplification of the isotherm model in the CLE–AdCSV method [53], it remains the only method for quantifying the DFe complexation situation. Over the past three decades, CLE–AdCSV has been the preferred method for studying trace metal speciation, including Lt concentrations and the K FeL ′ of the metal–unknown ligand complex [67].

2.2 Historical developments of HSs detection by electrochemical techniques

HSs are the hydrophobic fraction of natural organic matter that is retained on XAD-8 resin and eluted by alkaline extraction [10]. This includes fulvic acid (FA), a lighter and soluble fraction obtained by centrifuging at pH 1, and humic acid (HA), the remaining part. Several methods can measure HSs in natural waters, such as electrochemical techniques [68,69], chemiluminescence [70], spectroscopy [21], and elemental analysis [71]. However, due to their variable composition and the lack of a convenient analytical method, quantifying HSs in natural waters poses challenges in terms of understanding their influence on DFe in natural water. The use of electrochemical techniques for HSs detection can provide concentration data, which has been verified with UV–vis and fluorimetric methods using the International Humic Substances Society (IHSS) Suwannee River FA [22,69].

The advancement of methods for determining electroactive HSs (eHS) has followed the development of trace metal detection techniques. They are based on the complexes with molybdenum (Mo(vi)), Fe(iii), or Cu(ii) with eHS in samples, which can be adsorbed onto a mercury drop electrode to cause an electrochemical reaction [69,72,73]. The development process of determining the eHS method is presented in Table 1 and Figure 2. Beginning in 1986, Quentel et al. originally found the reduction wave of eHS-Mo(vi) [68] and developed a method to measure eHSs or refractory organic substances by complexing with Mo(vi), and it was further optimized in subsequent studies [22,72,74], as displayed in Table 1. They tried two voltammetry methods (Square-wave versus differential pulse modes), optimized measuring conditions, and studied the interferences from other metals or organic matters [22,72,74]. It was observed that high concentrations of polysaccharides could interfere with the reduction wave of eHS, which would overestimate eHS concentrations [22,74].

Research on HSs complexation with Fe was promoted by the Fe–HS method developed by Laglera and van den Berg and Laglera et al., which was catalyzed in the presence of bromate and increase the sensitivity of Fe–HS, which can measure FA and HA in natural waters including seawater directly [9,69]. The method of determining eHS complexed with DFe can offer a more relevant measure and more suitable for the Fe complexation capacity of eHS than measuring eHS at pH 2, as it can be conducted at the natural pH of seawater [22,69]. Laglera et al. confirmed that the complex stability and sensitivity of the electroactive Fe-SRFA species remain unaffected by interference from other metals and iron ligands in UV-digested seawater [69]. Based on the Fe–HS method [69], Bundy et al. and Yang et al. refined experimental parameters to study regional coastal waters [62,75]. However, the presence of a potent oxidant, such as bromate, used for catalytic reoxidation, can disrupt complexation with HSs during measurement. To address this, Laglera et al. and Sukekava et al. rearranged and expanded the initial analytical protocol [69,76]. This test method involved first collecting the natural Fe–HS complex single and then commencing the voltammetric analysis under iron saturation by the quantification of HSs standard addition that has been prepared in ultrapure water is carefully saturated with iron. This eHS method [69,76] has found extensive application in freshwater, estuarine, coastal waters, and the open ocean [25,77,78,79,80].

Cu–HS as the complex was also used to measure HSs concentration [73,81], which has been applied to study Cu and its speciation in estuary and coastal waters [82]. The interference and competition from HSs and other organic matters complexing with Fe, Cu, etc., have also been studied [73,81,83]. In the most recently updated method, the HSs in natural waters can be determined using a new, simple, and sensitive method based on their influence on the background current in a differential pulse-AdCSV, developed by Marcinek et al. [18]. The method was well correlated with classical spectrophotometry and even showed slightly better sensitivity than absorbance measurements [18]. The Mo–HS and Fe–HS methods were verified as compelling demonstrations of equivalent results by Dulaquais et al. [84]. Furthermore, Fe-binding and Cu-binding HSs in estuarine waters are the same [81]. Those HS methods have been applied to various regions to elucidate the biogeochemistry of DFe in natural waters [8,25,80].

In the determination of HSs using electrochemical techniques, the choice of HS reference standards is crucial for accurately showing the concentration of eHS. This is primarily due to the use of standard addition for quantification, which represented the real meaning of the determined eHS concentrations [22]. While carbohydrates and protein-like substances do not interfere with the measurement of individual HA or FA, different types of HA or FA standards indeed showed various results. The eHS determination needs to underline the importance of choosing the most appropriate standard. Nowadays, the IHSS supplies a few of organic matters references from different regions, while FA and HA reference materials originating from the Suwannee River are often used for the purpose of HSs method calibration and as model compound for speciation studies since 2007 [69]. In most studies about natural waters including seawater, eHS concentrations were often expressed as equivalents of Suwannee River humic and fulvic acid (SRHA and SRFA). Due to no standard extracting from seawater, thus the determination of eHS has unified SRFA as the standard references in mostly electrochemical method studies.

Electrochemical techniques have raised awareness that HSs have a strong complexation ability with DFe, which is sufficient to affect the biogeochemical action of DFe in seawater. Relevant studies have found that there are differences in the complex properties of SRHA and SRFA with Fe. Laglera and van den Berg determined the stability of the complexation of Fe with SRFA and SRHA (log K Fe'HS ′ ) to be 10.6 and 11.1, respectively [9]. Their complexation stability is similar to that of natural organic ligands (log K FeLt ′ between 9.6 and 13.9) [16]. The Fe-binding capacities of SRHA and SRFA are 32 ± 2.2 nmol Fe (mg SRHA)⁻¹ and 16.7 ± 2.0 nmol Fe (mg SRFA)⁻¹, respectively [9], while the Cu-binding capacities of SRHA are 18 nmol Fe (mg SRHA)⁻¹ [81]. The coordination capacity of SRHA for Fe is almost twice that of FA, but the coordination capacity of SRHA for Cu is similar to that of SRFA for Fe [81].

The observation of similar reduction potentials for marine eHS and SRHA or SRFA is more strongly related to the metal being reduced or oxidized than the compounds bound to metals on the mercury drop, such as the different reduction potential of eHS complexing with Mo(vi), Fe(iii), and Cu(ii) (Table 1) [69,74,81]. Therefore, it does not mean the peak of DFe–Lt complexation was single eHS [83] or mixed organic compounds or even polysaccharides [74,85]. Furthermore, our present knowledge of Lt and HSs is indirect because these ligands are complex, unknown identities and have never been isolated, recovered, and characterized. Thus, it was useful to know the weak complexation of electrochemical methods in measuring Lt and HS, which would be beneficial for the development of the terminological issues. According to existing research, some methods for determining organic matters, such as spectroscopic, chromatographic, and mass spectrometry, have supplied a lot of information on Lt or HSs, which would help analyze Lt or HSs together with AdCSV.

In future research, the application of electrochemical methods to detect trace elements in seawater is likely to diverge in two primary directions (Figure 2): the first being the development of portable or online electrochemical sensors and biosensors designed to detect heavy metals [86], and the second focusing on the mechanisms associated with complexing trace metals with ligands and HSs using multiple methods (spectroscopic, chromatography, and mass spectrometry) in seawater [26].

3.1 Exploring the effect of HSs on DFe through electrochemical methods

Over the past decade, scientists have attempted to use electrochemical methods to determine HSs’ concentrations. It is generally believed that HSs may be an important ligand of DFe and that they have an important impact on its concentration, distribution, and circulation [9,16,20,57,87]. The author collected DFe and HS concentrations from different types of water bodies, as shown in Table 2, which exhibited similar decreasing trends from river water to estuaries, coastal seas, and oceans, like Fe speciation modeling by Hassler et al. [6] (Table 2; Figure 1).

HSs and DFe have similar distribution trends in the Irish Sea; thus, natural HSs may account for all ligands in offshore waters [9]. Bundy et al. reported a direct linear relationship between the HSs concentration and Fe and weak organic ligand concentrations, which play an important role in Fe distribution [62]. It was verified that HSs account for virtually all of Lt of Fe in estuarine waters [81].

For the complex water body of the estuary area in early classical studies, HSs and Fe were observed to mainly exhibit consistent behavior no matter how salinity changed [88], which attracted great attention to further studies. Recent studies have verified that HSs play an important role in stabilizing DFe in various estuaries, mainly through the combination of HSs binding to DFe and the distribution of HSs, such as in the Mersey River Estuary [9,89], Thurso Bay [90], Yangtze River Estuary [75], and a boreal river in Newfoundland, Canada [91]. Stolpe and Hassellöv found that the HSs removal ratio in estuarine areas affects the amount of Fe imported into the sea by rivers [92]. In the rich HSs in the estuary of Mercer Bay in northern Scotland, Batchelli et al. found that colloidal Fe was mainly complexed by HSs, further indicating that HSs play an important role in the distribution of Fe in nearshore waters [90]. In the process of freshwater mixing with seawater along a salinity gradient, HSs flocculation occurs and plays an important role in stabilizing a small amount of DFe [58,93]. In a study of Arctic estuaries with rich organic matter, it was found that HSs contributed greatly to complexation with Fe, but the complexation intensity was different from that of temperate estuaries [94].

The DFe-binding capacity of HSs can explain why they play an important role in stabilizing the concentration and controlling the distribution of Fe [9]. Yang et al. reported that the Fe-binding capacity of HSs decreases exponentially with increasing salinity [75], which may be due to changes in the structure of HSs with increasing salinity [95]. In offshore waters, such as the Irish Sea and San Francisco Bay, the DFe-binding capacity at individual stations is lower than that measured under laboratory conditions, which indicates that HSs, as excess ligands, can form complexes with most of the DFe [9,58]. Similarly, according to data on the DFe-binding capacity of HSs in the East China Sea, HSs in most regions were also excess ligands for complexed DFe, although the concentration of DFe in the Yangtze River estuary was excessively high due to land-based inputs [75]. Therefore, Fe–HS complexes serve as the primary source of terrestrial iron reaching ocean waters [7,75].

Although HSs have been confirmed to play a pivotal role in stabilizing DFe, recent studies have offered alternative explanations for the complexation of DFe in specific seawater conditions [80]. The correlation between DFe and iron ligand concentrations in coastal waters has consistently been observed [96,97]. However, discrepancies have been noted between DFe and ligands in shelf and slope upwelling areas. Unlike expected, eHS concentrations did not display decreasing horizontal gradients from land to ocean. Instead, eHS were found to be iron–humic complexes that play a crucial role in complexing with DFe in upwelled waters [80]. This raises the need for further research into the sources or sinks that influence the distribution of HSs, ligands, and DFe in upwelling waters and other unique regions.

In open oceans, over 99% of the Fe is complexed with organic ligands, which increase iron solubility and microbial availability in early studies [96,98]. HSs’ concentrations were found to co-vary with DFe in the North and Northwest Atlantic, Southern Ocean, and Northeast Pacific [87]. In the Lau Basin of the western tropical south, including areas near hydrothermal vents [99], a substantial fraction of DFe is complexed by humic-like ligands (mean of approximately 30%). It has been confirmed that eHS can solubilize a substantial portion of freshly formed Fe-oxyhydroxides; however, this ability decreases rapidly as the inorganic particles age. Using an ocean circulation model, Misumi et al. [100] simulated how the distribution of weak ligands in the ocean controls the concentration of DFe. The model results showed that HSs play an important role in the Fe cycle in the ocean, which is consistent with some field research results [100]. Recent research has found that eHS accounted for only 20 ± 13% of Lt in the mixed layer and 8 ± 6% in deep waters within the western tropical South Pacific Ocean [25].

In the upper water column of open oceans, HSs may constitute only a fraction of the ligand pool (<20%, Figure 1), which includes terrestrial humic material and is supplemented by microbes originating from generic cellular debris as marine detritus [6]. In some open ocean studies, the eHS, as determined by electrochemical methods, had no correlation with the humic-like substance measured via fluorescence. This highlights the disparity in the detection of HSs using the two methods, illustrated by studies conducted from the central Pacific Ocean to the Bering Sea. However, a strong positive correlation between the concentrations of eHS and chlorophyll a suggests their biological origins [8], implying that the sources of eHS and DFe may differ in open oceans.

The latest research indicates that the electrochemical method not only quantifies eHS concentrations but also sheds light on the role of HSs in the marine iron cycle. This is achieved by examining the ligand exchange in a ternary SRHS-Fe-DFOB (desferrioxamine B, a type of siderophore), leading to the hypothesis that DFe complexed with various ligands might operate on a “first come, first served” basis [7]. These electrochemical techniques are not limited to measuring HSs’ concentrations and the binding capacity of DFe; they also facilitate our understanding of the competition among different ligands for DFe. Consequently, electrochemical approaches offer insights into the impact of HSs on DFe in marine settings.

3.2 Exploring the effect of HSs on DFe through other techniques

HS from marine or terrestrial sources or DOM in cool temperate zones to subtropical and tropical wetland systems have various compositions and structures, which may affect the effect of humic-like substances on iron in different rivers [103]. Spectroscopic, chromatographic, and mass spectrometry, as well as their hyphenated techniques, have been widely used to distinguish the structures, properties, and types of Lt and HSs [104,105,106]. Those methods can help us to know Lt and HSs from marine or terrestrial sources and can be used to trace the influence of terrestrial materials on seawater.

Therefore, an increasing number of scientists have focused on the role of humic-like substances and ligands in the Fe cycle of marine environments [20,86,107,108,109]. Many field studies have focused on humic-like fluorescent dissolved organic matter (FDOM) using optical methods [106,110]. In the nearshore area of the Yukon Estuary, humic-like FDOM was found to be correlated with DFe and play an important role in the transportation of DFe complexed with natural organic ligands to the continental shelf of the Bering Sea [111]. In the central sea area of the North Pacific, the DFe concentrations are strongly and linearly correlated with the fluorescence intensity of humic-like FDOM [112,113]. In recent spectral studies, the influence of humic-like FDOM not only occurred from the river plume to seas but also to open oceans. For example, terrigenous HSs are a regulator of the concentrations of DFe from the Gaoping River plume to seas [20]; marginal seas have been found to be important external sources of DFe in the North Pacific through the circulation of intermediate water, according to humic-like FDOM over the North Pacific [106]; DFe correlated with the humic-like FDOM in the western tropical South Pacific Ocean [25]; vascular plant and algal derived humic-like CDOM play a contributed role on DFe-ligand pool in the Laptev Sea and major Arctic basins [114].

Humic-like ligands are ubiquitous in the Lau Basin of the western tropical South Pacific Ocean and organic Arctic boreal soils, as well as proximate to hydrothermal vents, underpin the bioavailability and stabilization of DFe in oceans [99,108]. Sakata et al. found that humic-like substances in aerosol particles and surface water could control Fe solubility in size-fractionated aerosol particles in the Pacific Ocean using X-ray absorption–fine structure spectroscopy [115]. The latest laboratory experiments have provided new evidence of the importance of HSs in preserving DFe in estuaries and tidal flat wetlands [116,117]. Small aquatic humic ligands produced by organic soils in the Arctic boreal region also play a critical role in relieving the Fe limitation of phytoplankton in rivers and oceans [108]. Heerah and Reader identified the Fe-binding ligands present in a boreal river in Canada using Fourier-transform infrared spectroscopy, which verified the importance of the influence of humic-like ligands on the remaining DFe along an artificial salinity gradient [91]. Sato et al. found high amounts of humic-like substances in low-salinity surface waters between the outer edge of the Changjiang diluted water plume and a branch of the Kuroshio Current, which may have been influenced by microbial activity or photolytic degradation [118].

Except for exploring the influence of humic-like FDOM on the distribution of DFe in natural water samples, it was necessary to identify some kinds of matters of Lt and HSs and study interactions of DFe with them by chromatographic and mass spectrometry, as well as their hyphenated techniques at a molecular level. Waska et al. have found that the equilibration of DFe and artificial ligands can be investigated by electrospray-ionization mass spectrometry (ESI-MS) and Fourier-transform ion cyclotron resonance mass spectrometry (FT-ICR-MS), which may be as a representative tool to study interactions between trace metals and DOM [119,120]. Some kinds of chromatographic separation technologies, such as liquid chromatography (LC), immobilized-metal affinity (iMAC), and solid-phase extraction (SPE), would help ESI-MS and FT-ICR-MS supply more information about Lt complexing with DFe [105,121,122,123]. For example, siderophores in the eastern tropical North Pacific oxygen-deficient zone were exactly isolated and identified by LC-ESI-MS [104]; Gledhill et al. determined elemental stoichiometries of aluminum, iron, copper, nickel, zinc, cobalt, and manganese associated with a fraction of the DOM pool isolated by SPE-LC coupled to inductively coupled plasma mass (ICP-MS) [124]. Furthermore, isotope exchange LC−ICP-MS supplied us another technique for understanding Lt and HS, which can bridge molecular-level ligand identification with kinetic and thermodynamic metal-binding properties [19,125].

Therefore, the measurement of HSs by electrochemical methods, spectroscopic methods, chromatographic methods, and mass spectrometry, as well as their hyphenated techniques, have provided much information on the distribution and correlation of HSs complexed with DFe. In most estuaries, coastal waters, and open oceans, HSs or humic-like substances have a correlation with DFe; however, some special regions may have other influencing factors. This needs to explore deeply and widely on the source or sink, or other factors.