4.1. Soil Slurry Incubation
Our analyses detected significantly different dynamics and concentrations of Fe(III) reduction when compared between marsh and meadow wetlands. Using a soil slurry assay, we found that Fe(III) reduction in marsh wetlands responded more quickly and showed higher reductive activity (Figure 1
). There may be several factors controlling the rates of iron reduction in soils. Firstly, these two kinds of wetlands differ in terms of soil structure. The C. lasiocarpa
dominated marsh is characterized by a fibrous and brown root layer, a thin peat layer and a pale yellow and sticky gley soil layer in the soil profile. In contrast, in the D. angustifolia
dominated meadow, the soil is a typical meadow marsh soil classified as Inceptisols in the US soil taxonomy classification system [20
]. Therefore, soils particles, thermal properties, and the type and amount of substrates in soil are different (Table 1
). These differences may lead to differences in their respective drivers and the microhabitat of iron reduction. Peat soils tend to be finely textured, have high thermal properties and can transfer more heat to iron reduction microorganisms [21
]. Soil particle size and available surface area may also affect the bioavailability of Fe(III) oxides. Clay minerals could change the dissolved iron to a solid phase by absorbing charged Fe ions or colloids [22
]. The proportion of clay in soils (<0.002 mm) at 0–20 cm was 2.52% in the marsh, and was 39.32% in the meadow; this would lead to less dissolved iron for reduction in meadow wetlands. Furthermore, more substrates, such as organic acids, phenolic compounds, carbohydrates, and labile carbon, are expected to be used by bacteria in peatland soil than iron reduction in the marsh soils. However, Todoroval et al. found that there was a negative correlation between iron reduction and organic matter contents [21
]. The actual mechanisms underlying changes in organic matter in natural environments still need be determined.
Secondly, roots and rhizosphere microorganisms promote rapid enzymatic Fe(III) reduction in wetland soils [10
]. A previous study found that root morphology and chemical compositions, such as length, biomass, surface area, root carbon, and nitrogen concentration, were different when compared between C. lasiocarpa
and D. angustifolia
]. Furthermore, it has been suggested that Fe(III) mineralogy, carbon availability, microbial community, and Fe plaques (root-associated Fe(III) mineral deposits) in the rhizosphere is different when compared between these two wetlands. Roots are likely to represent a relatively abundant source of organic chelators as electron shuttles maintaining Fe(III) in a soluble form or humic compounds that can be used as electron donors, which can influence the rates of Fe(III) reduction. Additionally, the leaf and culm mass decomposition of C. lasiocarpa
and D. angustifolia
was different [24
], which may also affect Fe(III) reduction in these two wetlands.
Moreover, the concentration and phase of Fe(III) minerals are the main factors that control iron redox chemistry in soil environments. The frequent oxidation-reduction reactions in wetlands do not allow enough time for stable crystalline forms of Fe(III) oxides to form. Therefore, there is a tendency for amorphous and poorly crystalline forms of Fe(III) oxides to dominate wetlands [22
]. Poorly crystalline minerals of Fe(III), such as ferrihydrite, are thought to be more reducible than a more crystalline phase. Based on soil characteristics, the relative proportions of easily reducible Fe and amorphous Fe in marsh wetlands are much higher than that in meadows (Table 1
). The presence of more reducible Fe minerals leads to a greater content of ferrous iron. Therefore, a more significant microbial community of iron reduction, and a more suitable microhabitat, coupled with more available substrates in peatland, leads to a greater extent of Fe(III) reduction. To clarify the changes in soil microorganisms induced by organic carbon, research on microbial community abundance, and structures related to Fe(III) reduction should be further explored.
4.2. Microbial Inoculation Incubation
Dissimilatory Fe(III) reduction can be regulated by the addition of organic carbon. However, different compositions of organic carbon can affect degradation rates and related microbial communities, thus leading to differences in iron reduction activity. The addition of glucose and pyruvate can both promote the growth metabolism of iron reduction bacteria and rapidly accelerate the Fe(III) reduction process. In contrast, the lack of Fe(II) accumulation in our control treatment showed that an insufficient supply of electron donors could not stimulate Fe(III) reduction. Glucose had a higher capacity to transfer electrons to Fe(III) than pyruvate. While we were unable to find a similar study in wetlands, a similar conclusion was previously reported for paddy soils [25
In contrast to the case of pyruvate, the microbial metabolism of glucose can yield actual electron donors, such as H2
, pyruvate, acetate, and lactate. This can support Fe(III) reduction by providing preferential and abundant electron donors for Fe(III) reduction bacteria, and also for fermentative bacteria. Additionally, microbial dehydrogenation and hydrogen has been reported to couple with the fermentation of organic matter and play a significant role in microbial Fe(III) reduction [26
]. With an increase in nutrient source after glucose additions, the abundance of fermentation reducers, such as Clostridium, Pseudomonas, and Bacillus also increased. Clostridium was the main form of bacterial species present after the consumption of O2
and the main representative of fermentative microbes [28
]. Clostridium can produce a sufficient amount of H2
to act as an electron donor for the reduction of Fe(III).
Previous reports indicated that only a few dissimilatory iron reduction bacteria isolates were capable of using glucose directly [29
]. Hydrogen-dependent iron reducing bacteria are most likely to represent the main microbes at the beginning of flooding and contribute predominantly to iron reduction [30
]. Weber et al. [31
] also found that the community composition of microbes changed between the early and late stages of flooding. In the early stage, fermentative microbes tend to be dominant; while in the late stage, syntrophic acetate and H2
-utilizing methanogenic bacteria are more dominant. We should therefore pay more attention to separate and purify hydrogen-dependent iron reducing bacteria in the future.
Furthermore, the stable Fe(II) concentrations at the end of the incubation may indicate that intermediates transferred electrons to Fe(III) oxide, but not glucose directly. We also observed some bubbles in the glucose treatment, which may be a mixture of CO2
arising from the action of hydrogenase produced during glucose fermentation [32
AQDS can influence soil biogeochemistry not only in an indirect manner by changing soil structure and chemistry, but also by directly mediating the electron transfer process by functioning as an electron shuttle [33
]. AQDS as an electron shuttle between microbes and insoluble organic matter, can transfer more electrons to Fe(III), thus causing a distinct acceleration in the rate of Fe(III) reduction. In the presence of AQDS, complexation would render Fe(III) more accessible to micro-organisms. Our study showed that after the addition of AQDS, the Vmax was 30–50 times higher than in controls. Chen et al. [34
] also found that Fe(III) reduction bacteria were enriched by the additional presence of AQDS.
Moreover, AQDS, which has the ability to form humic-metal complexes, can also influence the amount of reduction as a terminal electron acceptor [11
]. However, the extremely high rates observed after the addition of AQDS in our study indicates that iron reduction was more limited by the availability of electron acceptors than by energy or mineral nutrients; this observation was also reported by Lipson et al. [6
] in an Arctic peat soil.
When there was no exogenous carbon source, the physiological metabolism of iron reduction stagnated, so the pH in the control treatment remained stable. This also indicated that the process of Fe(III) reduction relies upon enzyme reactions. The increase in pH was largely due to the consumption of H+. Thereafter, pH tended to stable or increased by only a small amount. This may have been caused by the short-chain fatty acids released during mineralization in floods. The pattern of pH generally reflected a negative relationship between H+ and iron reduction.