Seasonal antimony pollution caused by high mobility of antimony in sediments: In situ evidence and mechanical interpretation
Graphical abstract
Introduction
Antimony (Sb) is an important element which is widely used industrially, especially in production of flame retardants [1,2]. In recent years, Sb has gained increasing research attention due to its elevated concentrations in the environment and its toxicity [3,4]. Antimony and its related compounds can induce significant human toxicity, damaging the liver, heart and nervous system, as well as being carcinogenic [5,6]. Antimonites [Sb(III)] and antimonates [Sb(V)] are the prevalent oxidation states present in the natural environment [2,7]. Antimonates occurs in solution as Sb(OH)6− which is not known to form precipitates in the environment, unlike Sb(III) [8], while Sb(V) is strongly retained by metal (Fe, Mn, Al) oxyhydroxides under acidic conditions [[9], [10], [11]]. Yet the toxicity of Sb(III) is higher than that of Sb(V) [12], therefore the determination of Sb speciation is crucial for better understanding of the mechanisms of Sb transformation, bioavailability and toxicity in water and sediments.
Antimony speciation in sediments is largely controlled by the pH and redox conditions [13]. Mitsunobu et al. [14] indicated that Sb(V) is the mainly species with redox potential (Eh) ranging from 360 to −140 mV at pH 8 in a soil-water interface. In addition, thermodynamic computations indicated that the transition of Sb(V) to Sb(III) occurs approximately when Fe(III) goes to Fe(II), the condition that differentiates suboxic from anoxic systems [15]. Studies have shown that Mn(IV) and Fe(III) oxides can rapidly oxidize Sb(III) into Sb(V) within a few days, followed by the strong adsorption of Sb(V) by Mn(IV) and Fe(III) oxides via forming inner-sphere complexes at the mineral surface and the adsorption of Fe(III) oxides far exceeds that of Mn (and Al) oxides [10,[16], [17], [18]]. In addition, the oxidation of Sb(III) by Fe oxides is strongly depend on pH (favored at 3–5.9 pH) [10]. In contrast, in anoxic sediments microbial reduction processes can transform Sb(V) to Sb(III) [2,19]. Other factors, such as natural organic matter (NOM) and sulfide can directly influence the geochemical behavior of Sb in aquatic environments via complexation and precipitation reactions. Studies have shown that NOM are important organic ligands for Sb(III) binding [20], and the percentage of Sb(III) bound to NOM under environmentally relevant conditions is up to 30% [21]. An EXAFS study also revealed the presence of a Sb(III)-sulfur phase like amorphous antimony sulfide and antimony complex combined with organic sulfur functional groups in a contaminated freshwater wetland sediment [22]. It has been established that the mobility of Sb in sediments and water bodies is largely determined by reduction, precipitation and adsorption reactions [3,23], however mechanisms of Sb mobilization in sediments at high-resolution scale is largely unknown.
With rapid development and urbanization, anthropogenic eutrophication has become a major global environmental problem. Eutrophication stimulates algal bloom development, which in turn significantly changes the chemical and biological conditions of water and sediments [24,25]. Laboratory experiments studied the influence of algae on Sb speciation in the aquatic environment and found that algal blooms can promote the photochemical oxidation of Sb(III) to Sb(V), increasing the solubility of Sb [26]. Excessive algal growth can absorb large quantities of Sb through electrostatic attraction by surface functional groups [27]. In addition, changes in the lake-bed environment induced by algal bloom development and degradation (such as pH, oxic-anoxic condition and microbial community) [28] may significantly modify the mobility of Sb in sediments and its impact on water quality. However, little research has been undertaken in the field and the geochemical behavior of Sb influenced by algal blooms in eutrophic lake environments remains unclear.
Sediments are highly heterogeneous and show notable variation at the micro-scale, especially at the sediment-water interface (SWI) [29,30]. Therefore, the in-situ and high spatial resolution measurement of solutes is essential to establish their geochemical behavior in sediments [31]. In-situ passive sampling techniques, such as high-resolution diffusive equilibration in thin films (DET), dialysis (HR-Peeper) and diffusive gradients in thin films techniques (DGT), have increasingly been used for this purpose [2,32,33] and several DGT variants, such as ZrO-Chelex DGT, have been developed for Sb measurement [34]. The HR-Peeper method measures the concentrations of solutes in pore water based on diffusion equilibration [35], while DGT is a dynamic technique where measurement relies on diffusive gradients of the solute across a diffusion layer in the DGT device [36]. Instantaneous depletion of solutes in pore water at the surface of the DGT device are induced under DGT perturbance, thus leading to a further resupply of solutes from sediment solids to pore water [37]. Combination of HR-Peeper and DGT techniques has been increasingly used for the analysis of solute distribution and resupply kinetics in sediments [24,25,38,39].
In this study, a eutrophic bay of Lake Taihu (China) was selected, with a full-year investigation performed into Sb mobility in sediments and the impact on water quality. Monthly analysis of soluble Sb and labile Sb in the sediment-overlying water profiles was performed using HR-Peeper and ZrO-Chelex DGT, respectively. An aerobic-anaerobic experiment and an algal bloom simulation experiment were carried out to establish the mechanisms responsible for Sb mobilization in sediments.
Section snippets
Sampling sites
Lake Taihu, which is in the southern edge of the Yangtze River delta, is one of the five largest fresh water lakes in China. Meiliang Bay (31°26′18″ N, 120°11′12″ E) is in the northern part of Lake Taihu. It has 100 km2 surface area and 1.8–2.3 m depth. Meiliang Bay is the main water source for Wuxi City which is typically an algae-dominated region. In recent years, heavy metal pollution has been reported in Meiliang Bay at a moderate pollution level [40].
Monthly samplings were performed in
Monthly distributions of soluble and DGT-labile Sb
The vertical distributions of soluble Sb in pore water and DGT-labile Sb in sediments are shown in Fig. 1, Fig. 2. Soluble Sb and DGT-labile Sb concentrations mostly displayed small fluctuations in the vertical profiles in all months. Only soluble Sb concentration in December showed a considerable variation, reflected by high values (10.70–13.60 μg/L) in overlying water and upper 20 mm sediment layer followed by a decreasing trend with depth. The mean concentrations of soluble Sb and DGT-labile
Overall assessment of Sb pollution
Previous studies have focused on Sb cycling in aquatic systems through investigation of the changes in concentration of soluble Sb in the water column, as well as total concentrations and speciation of Sb in the sediment solid phase [[44], [45], [46]]. Due to the lack of in-situ information about the mobility of Sb at the sediment-water interface (SWI), the mobilization of Sb from the sediment and the associated potential Sb pollution risk in the overlying water remain unclear in the
Acknowledgments
This research work was financially supported by the National Natural Science Foundation of China (41571465, 41621002 and 41701570), CAS Interdisciplinary, Innovation Team, and Research instrument and equipment, and Development Project of the Chinese Academy of Sciences (YJKYYQ20170016).
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