SOC, MBC and soil respiration in PSFs under heavy metal pollution
A number of studies have shown that agricultural land management practices, such as crop planting, tillage and the irrigation state of the field, could affect the formation of aggregates via altering soil physical, chemical and biological properties (Kandeler & Murer, 1993; Six et al., 2000; Jiang et al., 2011; Gude et al., 2012). Such practices are also important drivers of microbial community dynamics and soil chemical processes. However, based on the objectives of this study, we chose to focus only on the effect of particle size under heavy metal pollution in field conditions. Because it is difficult to find two or more representative field plots under a similar state of long-term contamination, we finally selected two field plots under consistent crop cultivation and management practices as research sites, one of which had been polluted for c. 40 years, while the other was nonpolluted. Soil composite sampling was performed before the plots were ploughed and flooded for rice transplantation to minimise the influence of tillage. We acknowledge that although composite sampling likely underestimates the variability between the plots, it can also reduce heterogeneity that may exist in the field, and this methodology is commonly adopted and well accepted, particularly in ecological investigations of, for example, soil, water and sediment samples (Pennanen et al., 1996; Sekiguchi et al., 2002; Chiu et al., 2006; Farkas et al., 2007; Liu et al., 2012). Despite the limitations of the applied sampling design, the findings of the present study shed light on the impacts of variations in heavy metal pollution concerning decreasing microbial activity and diversity between PSFs.
Heavy metals, and particularly their bioavailable concentrations, can potentially affect soil quality-related parameters such as enzyme activities or related processes. Although decreases in MBC and in the culturable microbial population size from homogenised bulk soil samples under long-term heavy metal pollution are well documented in the literature (Vásquez-Murrieta et al., 2006; Li et al., 2009; Liu et al., 2012), the information regarding microbial populations at the microenvironment scale under pollution is scant. Separation of soil particles provides insight into the association between the distribution of organic C, heavy metals and microbial biomass and activity at a relatively small scale. In the present study, the total heavy metal contents were found to be markedly higher in both polluted bulk soils and PSFs than in the background soils. However, the contents of bioavailable heavy metals showed few differences between the PSFs, which may indicate an equal toxicity to microorganisms in these fractions. Nevertheless, heavy metal pollution resulted in a significantly lower MBC contents and soil respiration across the PSFs, regardless of the SOC quantity they contained. Significantly negative correlations between total and bioavailable heavy metals and the respiration, MBC contents and dehydrogenase activity were also observed, which indicated the effects of heavy metal pollution on the biological characteristics across the PSFs. Our results were in agreement with the findings of Li et al. (2009), who observed decreases in microbial biomass, enzymes and function under long-term acid metal stress.
At the microenvironment scale, coarse sand fractions are characterised by high concentrations of labile C and N originating predominately from plant residues, whereas silt and clay fractions are usually characterised by high concentrations of relatively stable organic C and N (Elliott, 1986; Six et al., 2000). Moreover, high SOC contents are usually found in clay and silt fractions, whereas low contents are observed in the sand fractions of most soils, such as in Calcaric Phaeozems (Kandeler et al., 2000), Calcic Chernozems and Cambisols (Stemmer et al., 1998) and Humic Dystrudepts (Chiu et al., 2006). In the present study, however, we detected the highest SOC concentrations in the coarse sand fractions, followed by the clay fractions, which generally agrees with our previous findings in rice paddy soils (Zhang et al., 2007; Zheng et al., 2007). Indeed, a similarly higher SOC content in the coarse sand fraction was also noted under rape/rice rotation in a Hydragric Anthrosol in which crop residues were returned to the soil (Jiang et al., 2011). It has been demonstrated that the organic C (fresh or labile) derived from crop residues is first incorporated into the coarse sand fraction during the initial decomposition period and subsequently accumulates and becomes stable in silt or clay soils (Angers et al., 1997; Six et al., 2000). Consequently, it is possible that higher fresh SOC contents could first accumulate in the coarse sand fractions, particularly in soils receiving large amounts of crop residues. In the present study, according to the observed SOC content and PSFs distribution (%) in the coarse sand fraction, we detected a large contribution of coarse sand to the SOC content of the bulk soil. Another reason for the higher SOC content observed in the coarse sand than in the clay fraction may be the high contribution of finer minerals (i.e. particulate organic matter) and clay to the coarse sand fraction during the formation of macroaggregates, as microaggregates are formed within macroaggregates (Six et al., 2000).
The variations in microbial biomass and activity observed between particles of different sizes under heavy metal pollution could likely be derived from differences in their available SOC content. The higher bioavailable SOC content in the coarse sand fraction could lead to increases in MBC because the bioavailable SOC derived from fresh residue is an important C source for microbial activity (Six et al., 2000). Singh & Singh (1995) reported that MBC contents were higher in the coarse sand fraction than in the clay fraction across forest, savannah and crop soils. However, the coarse sand fractions subjected to heavy metal pollution showed greater decreases in MBC and soil basal respiration than the clay fractions (Table 4), which could be attributed to the significant reduction in both bacterial and fungal gene copy numbers in the former fractions. This type of response also suggested a persistent toxicity or stress effect on soil microbial activity associated with long-term field conditions. The sharply decreased dehydrogenase activity observed in the coarse and fine sand fractions under pollution further supported the inhibitory effect of heavy metals on microbial activity or regarding low SOC availability. Dehydrogenase is an intracellular enzyme that participates in oxidative phosphorylation in microorganisms. This enzyme has often been correlated with the availability of organic carbon in soils (Serra-Wittling et al., 1995) and is assumed to be linked to microbial respiratory processes (Insam, 2001). Thus, it is likely that the toxicity of metals has a much greater impact on soil respiration and MBC in the coarse sand fraction than in the clay fraction. In contrast, the smaller decrease in respiration observed in the clay fraction under pollution compared with nonpolluted conditions in the present study could be attributed to the lower bioavailability of organic C in this fraction. Previous studies have confirmed that the clay fraction contains higher contents of carbon compounds (e.g. polyethylene and lipids) that are recalcitrant to decomposition by microorganisms and are generally more stable (Spaccini et al., 2001; Chen & Chiu, 2003). The clay fraction from rice paddies has also been found to show higher contents of humus and humic acid than the other fractions (Ding et al., 2006). In a recent study, Davinic et al. (2012) further revealed that although the silt and clay fractions exhibit a high SOC content, it consists of older and more stable forms of C. The much lower dehydrogenase activity observed in the clay fraction therefore revealed a lower availability of C and microbial activity. Alternatively, microorganisms might be more resistant to environmental stress due to the accessibility of substrates and protective agents in the clay fraction (Sessitsch et al., 2001). We further found a significantly higher metabolic quotient across the PSFs under heavy metal pollution, which has frequently been reported in studies in response to heavy metal contamination (Brookes, 1995). Positive correlations were observed between the total Pb and Cu contents and the metabolic quotient, which indicated less efficient C utilisation or higher maintenance energy requirements under adverse conditions (Giller et al., 1998). In fact, the differences in respiration observed in response to heavy metal stress may indicate that there were shifts in microbial community diversity within the PSFs.
Bacterial and fungal abundance in the PSFs under heavy metal pollution
To determine the differences in microbial abundance that contributed to the decreased microbial biomass observed under pollution, a real-time PCR assay targeting the bacterial 16S rRNA genes and fungal 18S rRNA genes was applied. Compared with the nonpolluted soils, a lower bacterial abundance was only found in the coarse sand fraction, which indicated that heavy metal pollution had few effects on the bacterial populations within the PSFs. Fresh residues or relatively labile constituents that are abundant in the coarse sand fraction could be beneficial for the bacterial population (Denef et al., 2001). Moreover, the smaller pore size and high resistance to desiccation and protection against predation found in the clay fraction could be beneficial for bacterial inhabitation (Rutherford & Juma, 1992). Most groups of soil bacteria could survive well, and a few could even be promoted under pollution, as observed in the bacterial DGGE profiles, suggesting tolerance of the bacterial community to heavy metals. An unchanged bacterial abundance was also observed by Liu et al. (2012), who reported no differences in bacterial abundance at two sites among four rice paddies subjected to long-term heavy metal pollution.
Previous studies have demonstrated that fungi prefer to feed on decomposable SOC and predominantly colonise labile and light fractions (Frey et al., 2003; Simpson et al., 2004). Fungal hyphae are also considered to improve aggregate stability by binding particles with extracellular polysaccharides and enmeshing microaggregates to form macroaggregates (Haynes & Beare, 1997; Bronick & Lal, 2005). Thus, it is not surprising that fungal abundance was found to be highest in the coarse sand fraction and lowest in the clay fraction in the present study. A heterogeneous tolerance of microorganisms to heavy metals has been found between short-term laboratory studies and long-term field experiments, and this tolerance also varies between sites (Giller et al., 2009). As reviewed by Giller et al. (2009), there is a range of sensitivity to heavy metals among various types of microorganisms, and a threshold for sensitivity likely exists. In the present study, a significantly lower fungi abundance was observed across the PSFs under pollution compared with that of bacteria, with the exception of the fine sand fraction, suggesting that heavy metal pollution had more long-term effects on the fungal community. However, a few studies have found contrasting results, and positive effects of heavy metal addition on the fungal community have even been observed in short-term laboratory experiments (Frostegård et al., 1993; Rajapaksha et al., 2004). One explanation for these divergent results could be that compared with fungi, bacteria can likely be well adapted to chronic toxicity or stress in long-term-contaminated soils due to their wide substrate utilisation profile and high metabolic activity. Additionally, bacteria dominate in rice paddies and may compete with fungi for substrates, resulting in greater stress on the fungal community. The reduced fungal community diversity shown by the DGGE profiles in the present study further substantiated the effects of heavy metals on the fungi under long-term heavy metal stress. Decreases in the levels of fungal PLFAs and a lower tolerance of the fungal community to heavy metal pollution along two different gradients in Scandinavian coniferous forest soils were previously described by Pennanen et al. (1996), who attributed these changes to decreases in the numbers of ectomycorrhizal fungi under pollution. In addition, Liu et al. (2012) also reported a reduction in the culturable fungal population as well as fungal PLFAs and gene copy numbers in four rice paddies under long-term heavy metal pollution compared with background soils. Decreases in fungal population PLFAs were also observed by Kandeler et al. (2000), who found that fungal PLFAs contents were reduced in the coarse and fine sand fractions following the addition of heavy metals (Zn, Cu and Cd) to a Calcaric Phaeozems under field conditions after 10 years of exposure. In this study, a significantly negative correlation was observed between extractable Zn and the fungal abundance (Table 5), which may indicate that bioavailable Zn might have an effect on the fungal population. Our results, derived from c. 40 years of heavy metal pollution, further supported the notion that the microbial responses produced under long-term field conditions (chronic toxicity or stress) differed from those in short-term laboratory assays (acute toxicity or disturbance) (Giller et al., 2009).
As bacteria and fungi are two principal groups in rice paddy soils, a decrease in fungal abundance could contribute to the lower MBC contents observed under pollution compared with nonpolluted areas. Archaea may be as abundant as fungi in terms of their rRNA genes. However, due to the low contribution of this group to basal respiration under aerobic incubation, the archaeal community was not addressed in the present study. In addition, a significantly positive correlation between fungal abundance and basal respiration (r = 0.513, P < 0.01) was noted across the PSFs, which indicated that the lower fungal abundance detected under pollution could be responsible for the lower soil respiration compared with nonpolluted conditions, as C respired from fungi constituted a higher proportion of total CO2 (Six et al., 2006).
Bacterial and fungal community structures in PSFs under heavy metal pollution
In the literature, coarse sand fractions are reported to exhibit lower bacterial species richness and diversity (Kandeler et al., 2000; Poll et al., 2003; Sessitsch et al., 2001) or higher diversity (Marhan et al., 2007) or few differences (Jackson & Weeks, 2008) regarding the bacterial composition in PSFs. In the present study, a greater number of specific DGGE bands and a higher H′ diversity index (Fig. 3) were observed in the coarse sand fraction than in the others fractions, regardless of the existence of heavy metal pollution. One possible explanation for this finding is that the coarse sand fractions most likely provide multiple microhabitats and more labile SOC for bacterial groups. Alternatively, microaggregates may be formed within macroaggregates and then released upon breakdown of the macroaggregates (Six et al., 2000). The bacterial community could be driven more by shifts in the chemical composition than in the quantity of soil organic matter, as reported by Davinic et al. (2012), who found higher levels of diversity in macroaggregates than in the microaggregates via pyrosequencing. The majority of the band types observed in the present study (B1-B3, B11 and B13-B17), particularly the dominant bands, were shared by all of the fractions, indicating that the bacterial community structure was highly stable. Sequencing of these bands revealed that Acidobacteria, Betaproteobacteria, Gammaproteobacteria and Chloroflexi were the predominant groups (Table 6), and these genera are commonly found in rice paddies, as reported in previous studies by our group (Hussain et al., 2011; Chen et al., 2013). These findings indicate that these bacteria are widely distributed and survive well across all fractions and may play an important role in the turnover and stability of SOC. The cultured representatives of Chloroflexi, such as the Anaerolineae lineage, are a group of bacteria consisting of filamentous, slow-growing, anaerobic heterotrophs that decompose carbohydrates and amino acids (Yamada & Sekiguchi, 2009) and are found at high abundance in anoxic soils. These species are assumed to potentially act as degraders of relatively recalcitrant carbon compounds such as phenol (Fang et al., 2006) and 4-methylbenzoate (Wu et al., 2001), which may further contribute to their dominance in the silt and clay fractions, where there are lower labile SOC and oxygen concentrations. The additional bands (B4, B6, B9, B12 and B18) that were relatively enhanced in the coarse sand fraction were assigned to Acidobacteria and TM7, in agreement with the findings of Mummey & Stahl (2004), who observed that Acidobacteria predominated in macroaggregates and the outer fractions of microaggregates. Although a predominance of Acidobacteria has been associated with low-C soils (Fierer et al., 2007) and is assumed to be of ecological importance in rice paddies (Kielak et al., 2008), our understanding of their function is still limited due to the scarcity of cultivated species. Several species (i.e. B1-B4, belonging to Chlamydiae and Acidobacteria) were clearly inhibited under heavy metal pollution, whereas other species (B7 and B10, belonging to Chlamydiae and TM7) were enhanced, resulting in an unchanged bacterial diversity. This last result is in accord with the findings of Zhou et al. (2002), who reported that chromium contamination in high-organic-matter soils did not greatly reduce diversity due to the heterogeneity of microorganisms both spatially and in terms of carbon resources.
Contrary to the variation in bacterial diversity observed under pollution, the fungal community structure exhibited a different DGGE pattern, with a decreased diversity index being observed across the PSFs compared with the nonpolluted soils. The smaller number of fungal DGGE bands and decreased diversity provided more evidence of decreased fungal abundance under pollution. PCA of fungal 18S rRNA gene DGGE profiles revealed clear separation of the polluted PSFs compared with the nonpolluted fractions. Additionally, the significantly negative correlations detected between all of the heavy metals and the H′ fungal diversity index further supported the inhibitory effect of heavy metals on the fungal community, which is primarily responsible for the decomposition of organic matter in the coarse and fine sand fractions (Marhan et al., 2007). Ascomycota and Basidiomycota are important fungal groups in most soils (Carlile et al., 2001). Hussain et al. (2011) also reported that Ascomycota dominated the fungal community in the rhizosphere of rice plants. In the present study, Ascomycota and Basidiomycota were found to be predominant across all of the fractions, which suggested a ubiquity of these species and an important role in agroecosystems. Caesar-Tonthat (2002) noted that polysaccharides produced by Basidiomycetes are involved in soil aggregation and suggested that fungal-derived material plays an important role in the stability of the soil structure. Glomeromycota, which are known to form hyphae and improve aggregate stability, were ubiquitous among the fractions, although present in very small numbers in the clay fractions. Additionally, although the method described by Stemmer et al. (1998) is widely adopted for the separation of PSFs, it may not be completely appropriate for determination of the microbial community, especially for fungi. The combination of low-energy sonication, wet sieving and repeated centrifugation procedures might have caused the destruction of community assemblages and even cell structure. It may also not be completely suitable for the efficient collection of fungi, which exhibit a diameter of 3–8 μm (Killham, 1994), associated with particle sizes below 2 μm. Furthermore, despite its importance in studies of microbial ecology, the PCR-DGGE technique has some drawbacks that must be taken into account. A single DGGE band, despite being obtained after careful consideration, might not represent a single phylotype or species. Thus, new methods and techniques are required to better explore the relationships between changes in soil organic dynamics and the microbiological status.