Introduction

Heavy metals (HMs) are often deposited in sediments due to metal hydrolysis, adsorption, and accumulation (Geng et al. 2022). Although a small portion of HMs is remained dissolved in the water column. As a result, sediments in aquatic media can either retain metals or move them into the water column via various mechanisms (El Nemr et al. 2007; Bastami et al. 2015). Because of their toxicity, non-biodegradability/persistence, and ability to bioaccumulate in the food chain, HMs constitute a significant hazard to the aquatic environment (Gopinath et al. 2010; Zhang et al. 2017). Metals have different environmental behaviors depending on their chemical speciation, as various fractions of HMs are critical to assessing their mobility and bioavailability (Nagajyoti et al. 2010). Sediments' HMs may return to the aquatic environment through biological activity or changes in temperature, salinity, redox potential (Eh), and pH (Yang et al. 2012; Zhou et al. 2020; Madadi et al. 2022). Various indices have been implemented in recent decades to determine the ecological risk or contamination severity of HMs in sediments (Yang et al. 2012).

The redox reaction is a vital quality in sediment and soil systems. For instance, observation of Woolston Canal sediments showed an acceleration in the dissolved concentrations of Zn, Ni, Cu, Pb, and Cr under oxidation conditions, albeit their concentrations were reduced under redox conditions (Hartley and Dickinson 2010). In wetlands, soil qualities change from aerobic and oxidative conditions on the surface to anaerobic and reductive conditions at the sub-layers (Caple 1996). At different depths, the structure of the sediments is governed by the water level, pH, temperature, and oxidation potential (Bundt et al. 2001; Agnelli et al. 2004). The Eh–pH diagrams can estimate the solubility of different chemical species in aqueous environments (Karbassi and Heidari 2015). Oxidation–reduction reactions regulate several biochemical processes in coastal wetlands (Zhao et al. 2020).

Geochemists quantify total metals involved in petrogenesis and their remobilization potential during early diagenesis or resuspension of deposited sediments using several sequential extraction techniques, i.e., Tessier Five-step chemical separation (Yoo et al. 2013; Ranjbar Jafarabadi et al. 2018; Madadi et al. 2021). The sequential partitioning of HMs in Pacific pelagic sediments could also determine the source rock of the sediments(Chakraborty et al. 2016). Sediments have five different fractions of HMs, including F1 (soluble and exchangeable), F2 (carbonate-bound), F3 (bound to Fe–Mn oxides, reducible), F4 (bound to organic matter and sulfides, oxidizable), and F5 (residual fractions) (Zhang et al. 2017). F1 may be immediately absorbed by biota, whereas F2, F3, and F4 may be converted to active forms by changing the sediment environment, i.e., redox conditions, pH, salinity, and total organic matter (Li et al. 2016). Nevertheless, F5 is inactive due to its attachment to the crystalline lattices of sediments; upon the arrival of anthropogenic HMs, they mainly accumulate in F1, whereas lithogenic metals accumulate in F5 (Palleiro et al. 2016).

Up to now, a few studies have addressed the effect of Eh–pH on the behavior of HMs in sediments. Besides, the mobility and bioavailability of HMs in Gomishan Wetland sediments under dynamic redox conditions have remained unexplored. Therefore, this study was primarily aimed to evaluate the quality of water and sediments via pH, Eh, electrical conductivity (EC), dissolved oxygen (DO), and chemical oxygen demand (COD). In addition, we assessed the effects of oxidation conditions and Eh-dependent factors on the release and association of HMs (Co, Cd, Cr, V, Mo, and Pb) with different sedimentary phases. Finally, the geochemical factions/sources of the sediments' HMs were identified using Tessier sequential extraction procedure.

Materials and methods

Sampling

Sediment and water samples were taken from the Gomishan Wetland in Golestan Province, northern Iran, on November 30th, 2020. About 20 kg of sediment samples were collected using Peterson grabs. 60 L of the wetland water was also collected. Water and sediment samples were transferred to the laboratory in previously N2 purified PET containers and stored at 4 °C.

Electro-photocatalytic decomposition experiment

There has been increasing interest in photo-assisted electrochemical process (PEC) for the degradation of organic matter. This method utilizes photocatalysis and electrocatalysis in synergistic cooperation resulting in enhanced degradation performance (Yu et al. 2021). A photoanode (Ti/TiO2) and cathode (graphite) were used for the electro-photocatalysis tests with the dimensions of 1 × 20 × 100 mm and 10 × 50 × 100 mm, respectively (BaojiQi Xin dioxide Titanium Co., Ltd). The electrodes were placed in the aquarium at a distance of 7 cm in a parallel arrangement. Then, a DC-regulated power supply (UTP3704S UNI-T) was used with five positive and negative outputs. Anode was connected to the positive and the cathode was connected to the negative terminal to apply the voltages of 5 and 7 V. The UV lamp (8 W power; wavelength 380 nm; Philips Actinic BL TL) was set at a distance of 3 cm from the anode. Figure 1 schematically illustrates the PEC reactor system.

Fig. 1
figure 1

Schematic diagram of the PEC reactor system

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Before each test, the electrodes were rinsed with distilled water and 90% alcohol and oven-dried at 110 °C. To simulate the wetland condition, approximately 1 cm of sediments and 7 cm of water were added to each aquarium illuminated with a 250 W tungsten lamp (visible irradiation). The experiments were carried out at room temperature (25 °C), although the solution temperature increased by 10 °C after the PEC process due to the heat generated by the UV lamp.

The experiments were conducted under five experimental conditions, including 5 V, 7 V, 7 V/lamp/UV, 7 V/lamp, and 7 V/UV. Two electrodes were set at a 7 cm distance in all experiments.

Analytical methods

Temperature (T) was measured by a mercury thermometer; a portable multi-parameter probe was used to measure pH, Eh, DO, and EC. These measurements were all in situ. COD was determined using a spectrophotometer (DR 5000-Hach). Inductively coupled plasma mass spectrometry (ICP-MS) was employed to measure the concentrations of Co, Cd, Cr, V, Mo, and Pb (PerkinElmer, America). To calculate the concentrations of HMs and COD, 400 cc of water (100 cc for HMs and 300 cc for COD) was collected from each aquarium after 6 h and filtered through a 0.45 µm Whatman filter. To reduce and stabilize the solution pH < 2, H2SO4 (98%) and HNO3 (65%) were added to COD and HMs samples, respectively.

Tessier sequential extraction method (Tessier et al. 1979) was applied to determine the different fractions of HMs as follows: exchangeable (F1), acid-soluble (carbonates) (F2), reducible (Fe–Mn oxide bound) (F3), oxidizable (organically bound + sulphide bound) (F4), and residual (F5). The suspension was centrifuged at 3000 rpm for 15 min after each step to separate the obtained extract from the sediment. The supernatant was filtered and acidified with HNO3 to pH 2. The preceding phase's residue was washed with 10 mL of deionized water, waggled for 10 min, and centrifuged at 3000 rpm for 15 min, with the washing being thrown away in each step. Pre-dried sediment subsamples were digested using procedure 3050B to measure the residual and total concentrations of HMs. Finally, the concentration of HMs was measured by ICP-MS (PerkinElmer, America).

Statistical analysis

The present research used HSC Chemistry 6 software package. Inserting the Eh and pH values into HSC Chemistry 6 yields the predominant elements' temperature and concentration, chemical forms, and oxidation state in the aqueous system. Further statistical processing was made by making use of Excel 2019.

Sediment assessment

To determine the environmental pollution intensity, the geo-accumulation index (Igeo) was used as follows:

$$I_{{{\text{geo}}}} = {\text{Log}}_{2} \left[ {\frac{{{\text{Cn}}}}{{{\text{Bn}}_{x} 1.5}}} \right],$$
(1)

where Igeo refers to the soil pollution intensity index, Cn is the HM concentration (soil or sediment), Bn is the HM concentration in the shale, and 1.5 is the shale correction factor. Chemical partitioning can accurately determine Bn value through the separation of anthropogenic and natural sources of HMs.

Since the initial concentration of the element is not acquired through complete partitioning, the concentration of shale, a kind of sedimentary rock, should be utilized in this calculation. The average element concentration in shale is lower than the average element concentration in unpolluted sediments. Hence, it should be multiplied by 1.5 to achieve a balance. Based on this, (Karbassi et al. 2008) optimized the Müller formula as follows:

$$I_{{{\text{Poll}}}} = {\text{Log}}_{2} \left[ {\frac{{{\text{An}}}}{{{\text{Dn}}}}} \right],$$
(2)

where Ipoll is the pollution intensity, An is the HM concentration (soil or sediment), and Dn is the background concentration of the HM within the study area.

The accumulation factor for each element was calculated via the enrichment formula of Buat-Menard and Chesselet (1979) as follows:

$${\text{EF}} = \frac{{\frac{{{\text{Cn}}_{x} }}{{{\text{Bn}}_{x} }}}}{{\frac{{{\text{Cn}}_{y} }}{{{\text{Bn}}_{y} }}}},$$
(3)

where EF is the enrichment factor, Cnx is the concentration of sediment HM, Bnx is the Fe concentration of the sediment sample. Cny and Bny are the background concentrations of HM and Fe, respectively. This method uses Fe as a normalizing element to subdue changes arising from sediment heterogeneity (Abrahim and Parker 2008).The pollution index and the enrichment factor classes are shown in Table S1.

Water assessment

In this study, the heavy metal evaluation index (HEI) was used to assess the extent of HMs' pollution and water quality. The HEI index gives the total water quality with regard to the HMs using the following equation:

$${\text{HEI}} = \mathop \sum \limits_{i = 1}^{n} \frac{{H_{{\text{c}}} }}{{H_{{{\text{mac}}}} }},$$
(4)

where Hc and Hmac represent the measured and the maximum permissible concentrations of the ith parameter, respectively (Prasanna et al. 2012). The saltwater quality requirements for aquatic life were used to calculate Hmac values.

Results and discussion

Physicochemical properties of water samples

Table 1 shows the mean values of the physicochemical properties of water samples.

Table 1 Introductory analysis and parameters of water control samples taken from the Gomishan Wetland
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Shrimp breeding is the dominant farming practice in the sampling area and this can be addressed as the main reason for the relatively high salinity (12.77 PSU) and alkalinity of Gomishan Wetland water samples. Basins with high agricultural or urban intensities have higher salinity and pH than the adjacent areas with different land use (Chessman and Townsend 2010). The Eh mean value revealed an aerobic condition in the studied wetland (Goncharuk et al. 2010).

Heavy metals concentrations

Table 2 shows the mean heavy metal concentrations in surface sediment samples. The mean contents of Cd, Cr, Co, Pb, Ni, Zn, V, and Mn were 0.2,36,5.5,5,23,44,53, and 357 mg·kg−1, respectively. The mean concentrations of Cd, Co, Pb, V, and Zn in Gomishan Wetland sediments were lower than in similar studies. The concentration of Mn in the current study was significantly higher than previously published levels for Queensland (Qiu et al. 2011). Cr and Ni also showed relatively lower concentrations in comparison with similar studies (see Table 2). Based on Table 2, the mean metal concentrations in this study are lower than those in the earth's crust. Overall, it is possible to conclude that the amount of studied metals in the surface sediments of the Gomishan Wetland is almost less than that of the other studies mentioned in Table 2.

Table 2 Mean concentration of metals in some wetland sediments around the world (mg/kg)
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The effect of voltage variations on the PEC process

Table 3 presents the effects of the PEC processes on pH, Eh, EC, and DO under different conditions. pH, Eh, EC, and DO ranged between 7.20–7.42, 208–487 mV, 4.70–6.24 mS/cm, and 8.30–9.88 mg/l, respectively.

Eh, DO and COD degradation increased with the increasing voltage in samples. This can be justified, because oxygen is an electron acceptor and plays an essential role in separating electron–hole pairs. Oxygen reacts with two electrons and two hydrogen atoms, reducing up to H2O2. The latter produces a hydroxide radical. Accordingly, Oxygen increases the oxidation rate and COD degradation(Wang et al. 2017). The oxidation efficiency could result from rising the redox potential (Ebraheim et al. 2021).

The photocatalytic process decreased COD values significantly by 82% from 1986 mg/l for a blank sample to 350 mg/l at the base potential of 7 V (Table 3), which is consistent with the results of (Upadhyay et al. 2019) and (Fernandes et al. 2019). Our results also demonstrated that the increase of anodic potential intensified the COD reduction, which is consistent with the findings of (Chimonyo et al. 2017). These authors found a linear relationship between redox potential and the rate of COD removal. The increase of the anodic potential accelerated the movement of electrons, separation of electron–hole pairs, production of oxidant on the electrodes, and the oxidizing potential of PEC (Ghimire et al. 2019) (see Table 3).

Table 3 Post-PEC physical and chemical parameters of the samples
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pH change affects the PEC process through the metal oxide's surface charge, adjusting the pollutant's adsorption capacity on the surface, electron–hole separation, and oxidation potential of the valence band (Tayyebi et al. 2019). As depicted in Table 3, after oxidation, pH decreased compared with its initial value (7.9), ranging from 7.20 to 7.42. Eh and EC ranged from 487 to 208 (mV) and 4.70 to 6.42 (mS/cm), respectively, indicating a linear relationship between pH and EC and a nonlinear relationship between pH and Eh (Ustaoğlu and Tepe 2019). The decrease in solution pH increased with the mobility of HMs (Hartley and Dickinson 2010).

Heavy metals' mobility under different Eh and pH values

Figure 2 shows the release of selected HMs concerning Eh. As can be seen, the initial dissolved concentrations of Co, Cd, Cr, V, Mo, and Pb (Eh = 109 mV) were 3.2, < 1, 161.6, 44.9, 62.3, and 20 µg/l, respectively (Table 4). The increase in Eh value after PEC released the HMs from the sediment into the water (Fig. 2). This is in line with the result of (Hartley and Dickinson 2010). These authors found the increased mobility of trace metals and the decreased solution pH after drying and aeration of the canal sediment. However, our findings contradict the results of some others (Frémion et al. 2017; Ranjbar Jafarabadi et al. 2018; Madadi et al. 2021). The increase in mobility and release of metals from sediments can be due to the decrease in pH; the negative surface charge of colloids decreases with decreasing pH; thus, HMs desorb from organic matter and other sorbent surfaces, such as clay mineral surfaces (Shaheen and Rinklebe 2017). According to Fig. 2, the highest mobility of Co, Cd, Cr, and V occurred at Eh 369 mV with concentrations of 6.3, 2, 192.7, and 60.8 µg/l, respectively. The concentration of Pb increased after the oxidation, but it remained constant (40µg/l) at Eh values of 208, 235, and 487 mV. Mo had the highest dissolved concentrations (163.3 µg/l); before the Eh rise, it was 62.3 µg/L. The rise in the concentrations of dissolved HMs might attribute to the impact of Eh on metal species (Du Laing et al. 2009). The oxidation state of HMs species is a significant regulating factor in their behavior and bioavailability, since it directly affects the mobility, adsorption, and precipitation of HMs (Chakraborty et al. 2019).

Fig. 2
figure 2

Release of selected HMs by changing Eh

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Table 4 Heavy metal mean concentrations (µg/l) in water under different conditions
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Geochemical fractions of heavy metals

The chemical partitioning results showed that HMs exist in large quantities in the terrestrial phase (Fig. 3). Anthropogenic intrusions are responsible for the exchangeable and carbonate-bound portions. Contaminants from these fractions can quickly enter the water column, posing a severe threat to aquatic life (Saleem et al. 2018). The sulfide and organic matter (F4) were the dominant fractions of Cr (77.5%), V (77%), Mo and Co (76%), Pb (68.8%), and Cd (57%). A change in Eh might discharge the materials from F4 into the water, threatening water quality (Frémion et al. 2017). This reflects the metals' high affinity for sulfide and organic matter. Furthermore, the oxidizable fraction consists primarily of organic bind metals and sulfide-bind metals, with spatial differences in the main components of oxidizable fractions of different metals (Zhao et al. 2021). Metal components in F1, F2, F3, and F4 have a higher potential for mobilization from sediment and bioavailability to benthic organisms (Gao et al. 2010). Toxicity occurs when the rate of metal intake exceeds the rates of metal detoxification and excretion (Zaaboub et al. 2015).

Fig. 3
figure 3

Geochemical fractions of Cd, Pb, Co, Mo, V, and Cr in sediments

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The mean residual (F5) percentages of HMs’ in the Gomishan Wetland sediment were as follows: V (19.6%) > Cr (19.5%) > Co (19.3%) > Mo (19%) > Pb > (17%) > Cd (13.5%). In residual form, HMs are substantially less hazardous to aquatic organisms due to their chemical stability and biological inactivity (Wang et al. 2010). Mineralogy and weathering control the concentration of HMs in this phase (Shaila and Naotatsu 2010).

Indices of contamination

Pollution index (I poll)

The intensities of HMs' pollution index of Co, Cd, Cr, V, Mo, and Pb are shown in Table 5. To determine the contamination grades of HMs in the sediments of Gomishan Wetland, the findings were assessed based on the classification system presented in Table S1.

Ipoll values of HMs followed the order of Cd > Pb > Co > Mo > V > Cr before, and Cd > Pb > Mo > Co > Cr > V after oxidation. The Ipoll index values of all HMs were less than 0.5, indicating that the HMs were within the 'No pollution' range. Except for Cd, the amount of Ipoll of HMs decreased with increasing the Eh, which is consistent with the results of (Mohiuddin et al. 2012). According to their findings, Ipoll rose with decreasing Oxidation Reduction Potential (ORP) in winter as temperature decreased. Correspondingly, an increase in Eh from 109 to 487 mV increased Ipoll of Cd from 0.486 to 0.667, suggesting 'Low Cd pollution' (Table 5).

Table 5 Intensity of heavy metal pollution in Gomishan Wetland sediments
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Enrichment factor

Table 6 shows the severity of sediments pollution in Gomishan Wetland based on the enrichment factor (EF). The EF values suggested depletion to the minimal enrichment for Cd, V, Co, Pb, Cr (< 2), and significant enrichment for Mo (5 < EF = 6.88 < 20). Mining, fertilization, the manufacture of alloys, biosolids, catalysts, coal, and petroleum combustion are all causes of excessive Mo emissions into the environment (McGrath et al. 2010; Zaborowska et al. 2016). Based on this, we concluded that the agricultural activities around the study area are the main causes of high Mo concentration.

Table 6 Enrichment factor of heavy metals in Gomishan Wetland sediments
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Heavy metals evaluation index

Heavy metal evaluation index (HEI) calculates the mean concentrations of HMs in different water samples based on their concentration in aquatic water. The HEI values > 20 indicate poor water quality. In Gomishan Wetland, the minimum and maximum HEI values ​​were 27.23 and 40.25, respectively, occurring at Eh 208 and 487 mV. Figure 4 depicts the dispersion of HEI values. The initial HEI value was 38.59, and it reached 27.23 with rising Eh from 109 to 208 mV. Furthermore, the HEI values were 33.76, 33.11, 31.30, and 40.25 at Eh values of 235, 369, 381, and 487mV, respectively. Although the behavior of HMs in aquatic environment depends on many factors, Eh might be considered as a major one (Fig. 4). Changes in HEI values attribute to the adsorption–desorption of HMs by the sediments. Absorption of HMs into the sediment can decrease the HEI value, whereas the release of HMs from the sediment into the water environment can increase the HEI value (Madadi et al. 2021).

Fig. 4
figure 4

Changes in the values of HEI by increasing Eh

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HSC chemistry

The Eh–pH diagram displays the thermodynamic stability of HMs in the aqueous environment. The stability range often appears between two dotted lines. Values above this range indicate that the element is potentially toxic to the aqueous environment. Figure 5 reveals that Co, Cd, Cr, V, Mo, and Pb were all between the stability lines, posing no contamination or toxicity to the aqueous environment.

Fig. 5
figure 5

Specification of Cr, Cd, Co, Mo, Pb, and V

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Figure 5 shows that, at low solution pH (0–4), the main existing forms of HMs were Co+2, Cd+2, Cr+2, V+2, and Pb+2. The positive TiO2 surface found it difficult to absorb these cations. As a consequence, it is unable to produce sufficient electrons and OH. While the pH changed from 6 to 12, TiOH was created, and the TiO2 surface became negatively charged, allowing for further adsorption. Co +2, Cr+2, Mo+2, and Pb+2 ions were oxidized to Co3O4, Co(OH)2, MoO2, MoO4(-2), Cr2O3, CrO4(-2), Cr (OH)4(-1), PbO, PbO2, Pb12O17, Pb4 (OH)4 (+ 4), and Pb6(OH)8(+ 4), respectively, at solution pH between 6 and 12. The predominant species of Cr, Cd, Mo, V, and Co were Cr2O3, Cd +2 ions, MoO4−2, HV10O23−5, Co +2, and Co3O4 in the aquatic environment. Black and colored dots demonstrate the HMs' species before and after the PEC process, respectively.

Conclusion

The behavior of HMs under dynamic redox circumstances could be used to assess the potential risk, determine the destiny of HMs, and develop effective environmental management methods for the restoration of contaminated sites. Our findings indicated that an increase in solution Eh will augment the release of sediments' HMs into the water, which might contribute to the pH reduction. The highest mobility of Co, Cd, Cr, and V occurred at Eh 369 mV. The concentration of Pb increased in water after the oxidation, but it remained constant (40µg/l) at Eh values of 208, 235, and 487 mV. The highest release of Mo occurred at Eh 487 mV.

Sediment samples of Gomishan Wetland were analyzed to evaluate the contamination level of five heavy metals (Cd, Co, Cr, Mo, V and Pb). The findings showed that EF and Ipoll values of the studied wetland were within no pollution range, meaning this wetland receives medium levels of anthropogenic contamination compared with background values. Although for Mo, the EF value 5 < EF < 20, indicating anthropogenic sources. Moreover, Tessier sequential extraction method revealed that terrestrial fractions were the main sources of sediments' HMs. Future research should document the factors (sediment size, microbial activity, organic matter, temperature, and salinity) influencing the release of sediments' HMs. In addition, further studies should include the role of radical OH under various electrical conductivity (EC) ranges along with the other voltages. In addition, the kind of organic matters in sediments should be analyzed.