1- Introduction

2- Material and methods

3- Results

4- Discussion

5- Conclusion

References

Abstract

Soil microbes play vital roles in energy flow and nutrient cycling and, thus, are important for agricultural production. A better understanding of the complex responses of microbial communities to various organic and inorganic fertilization regimes is critical for sustainable development of agroecosystems. Changes in bacterial and fungal abundance and diversity in fluvo-aquic soil in Northern China were studied under 38-year long-term fertilization strategies: four chemical-fertilization strategies (i.e., no fertilizer, N, NP, or NPK), with or without manure amendment, were investigated by high-throughput sequencing and quantitative polymerase chain reaction-based amplification of bacterial 16S rRNA and fungal internal transcribed spacer (ITS) rRNA genes. Chemical fertilizer plus manure addition clearly increased the soil fertility and was recommended for further optimization of fertilization patterns. Both principal component analysis and partial least-square discriminant analysis showed greater impacts of manure addition than chemical fertilizer on bacterial community distributions, whereas fungal communities were more sensitive to inorganic fertilizer. The linear discriminant analysis effect size method revealed that the number of responding microbes (microbes significantly affected by various fertilizations) in bacterial communities in manure-treated soils was markedly higher than that in chemical fertilizer-treated soils, whereas those of fungal communities showed the opposite trend. In addition, redundancy analysis further illustrated the primary importance of organic matter in shaping community distributions of bacteria, rather than in driving fungal community patterns. These results suggested that organic and inorganic fertilizers, respectively, dominated in shaping bacterial and fungal community distributions in fluvo-aquic soils.

Introduction

Soil microorganisms play vital roles in the maintenance of soil health, productivity, and sustainability, and thus crop production, because of their contributions to soil nutrient transformation, formation and decomposition of organic matter, and stabilization of soil aggregates (Van Der Heijden et al., 2008; Zhao et al., 2014; Singh, 2016; Pan et al., 2018a). Therefore, a profound understanding of soil microbial ecology has become increasingly recognized and emphasized because of the ecological and economic importance of microorganisms (Mele and Crowley, 2008). Due to the increasing concerns regarding food demands and the scarcity of land available for agriculture, high amount of fertilizer is implemented to increase crop productivity (Inselsbacher et al., 2010). The application of organic and inorganic fertilizers can cause detectable changes in nutrient availability to plants, as well as in the diversity and function of microorganisms (Marschner et al., 2003). For example, long-term field experiments involving polymerase chain reaction (PCR)-denaturing gradient gel electrophoresis (DGGE) demonstrated the importance of balanced fertilization in promoting the activities of bacterial communities (Chu et al., 2007a,b; Luo et al., 2015). Nitrogen (N) is regarded as a key limiting factor for soil microbes, and N amendment can alter the activity and diversity of soil organisms (Sarathchandra et al., 2001; Xu et al., 2017; Liu et al., 2019). Organic fertilization exerted no effects on the fungal biomass, while increasing the bacterial biomass by increasing soil organic carbon (Corg) and C/N ratio (Crecchio et al., 2001; Marschner et al., 2003). In addition, manure application promoted soil bacterial communities by enhancing root exudates and also improved plant growth (Bittman et al., 2005). Suzuki et al.

Abstract

This research focuses on the heavy metal contamination of the paddy soils and rice from Kocˇani Field (eastern Macedonia) resulting from irrigation by riverine water impacted by past and present base-metal mining activities and acid mine drainage. Very high concentrations of As, Cd, Cu, Pb and Zn were found in the paddy soils (47.6, 6.4, 99, 983 and 1,245 lg g-1 ) and the rice (0.53, 0.31, 5.8, 0.5 and 67 lg g-1 ) in the western part of Kocˇani Field, close to the Zletovska River, which drains the mining facilities of the Pb–Zn mine in Zletovo. In terms of health risk, the observed highest concentrations of these elements in the rice could have an effect on human health and should be the subject of further investigations.

Introduction

Base-metal mining together with milling and grinding operations, concentrating ore and the disposal of tailings along with the acid mine and mill wastewater are some of the most important sources of heavy metals in the environment (Adriano 1986). Therefore, elevated concentrations of heavy metals can be found in and around abandoned and active mines due to the discharge and dispersion of mine waste materials, including tailings and acid mine drainage effluents into nearby agricultural soils, food crops, riverine water and stream sediments (Jung 2001; Korre et al. 2002; Lee et al. 2001; Liu et al. 2005; Lu and Zhang 2005; Simmons et al. 2005; Witte et al. 2004; Wong et al. 2002; Yang et al. 2004). It is well known that heavy metals in agricultural soils can influence both crop yield and quality, and soil containing various elements is a primary nutrient medium for plants. Food crops grown in contaminated areas may accumulate much higher than normal amounts of heavy metals, which may pose a significant public health risk to the local community.

The anaerobic degradation of organic matter accounts for a significant part of the global carbon cycle and involves the production of the greenhouse gas CH4 (Reeburgh, 2003). While the terminal processes liberating CH4 have been thoroughly investigated using isotopic techniques (Conrad, 2005), the contribution of other pathways during the anaerobic degradation is not well investigated. In this study we aimed to specify the contribution of acetogenesis to the acetate pool in rice paddy soils from two geographical origins (Philippines and Italy). The unifying trait of acetogenic bacteria is that acetate is their major product, which is mainly formed by the acetyl-CoA pathway reducing CO2 to acetate. However, the usage of the acetyl-CoA pathway in acetogens is versatile (Drake et al., 2002; Schuchmann and Muller, 2016): acetogens can compete with primary fermenters for monomeric compounds, e.g., glucose, fructose or xylose; they can act as secondary fermenters using lactate, ethanol or even methoxylated aromatics; many acetogens can grow chemolithotrophically converting H2 and CO2 to acetate; and some acetogens can even grow in syntrophic associations at the expense of acetate running the acetyl-CoA pathways in the reverse direction (Hattori et al., 2000; Hattori, 2008). Even though acetate is the major end product of acetogens, their environmental role is hard to address since many anaerobic bacteria also produce acetate by other biochemical pathways than the acetylCoA pathway. To investigate the function of chemolithotrophic acetogens in rice field soils we studied their contribution to the acetate pool using isotope techniques. This is a reasonable approach since studies of microbial pure cultures have shown that the stable carbon isotope fractionation of most fermentative processes is almost negligible (< 5‰) (Blair et al., 1985; Penning and Conrad, 2006; Botsch and Conrad, 2011), while in contrast the fractionation by chemolithotrophic acetogenesis (4 H2 + 2 CO2→ CH3COOH + 2 H2O) is strong (−38‰ to −68‰) (Gelwicks et al., 1989; Preuss et al., 1989; Blaser et al., 2013). One has to keep in mind that the acetate formed by heterotrophic acetogenesis originates partially from fermentation of organic matter and partially from the reduction of CO2 (Freude and Blaser, 2016). However, while the differentiation of aceticlastically and hydrogenotrophically formed CH4 is rather straight forward (Conrad, 2005; Blaser and Conrad, 2016), the application of such techniques for the investigation of chemolithotrophic acetogenic bacteria is much more complex. In contrast to CH4, which is an end product in most anaerobic sediments, acetate is produced and consumed by different metabolic pathways, which all imprint a signature on the stable carbon isotope composition (13C/12C) of acetate (Fig. 1) (Conrad et al., 2014). Hence, the stable carbon isotopic composition of acetate is basically governed by three factors: (1) the isotopic composition of its precursors, (2) the isotopic fractionations (enrichment factors) associated with its formation and consumption, and (3) the relative rates of all acetate producing and consuming processes which influence its pool size. The knowledge of all three factors is of great importance for the evaluation of 13C values of acetate determined in environmental samples.

Terrestrial ecosystems represent a major sink of CO2 through fixation by plants but they have been shown to mitigate the rise of atmospheric CO2 also via microbial CO2 fixation (Ge et al., 2016; Yuan et al., 2012). Microbial CO2 fixation has been mostly ascribed to autotrophic microorganisms (Ge et al., 2016), but fundamentally all microorganisms may use inorganic C (IC; i.e. CO2 or bicarbonate) in their metabolism. All these fixations require energy generated by phototrophic, autotrophic or heterotrophic energy sources. IC is the main or even the only C source for chemoautotrophs and photoautotrophs, while heterotrophs and mixotrophs rely on organic C (OC) but also incorporate IC via a variety of carboxylation reactions that are part of their central or peripheral metabolic pathways (for review see Erb, 2011; Wood and Stjernholm, 1962). The importance of carboxylases in heterotrophic metabolism increases whenever microorganisms experience C limitation through a disproportion between C demand for energy generation and growth and its availability, caused by deficiency or complexity of OC sources, or fast growth (Alonso-Saez et al., 2010; Feisthauer et al., 2008; Merlin et al., 2003). Even though the occurrence of dark and largely heterotrophic CO2 fixation in soils is generally accepted, very few studies have assessed its relevance for soil microorganisms (Miltner et al., 2004, 2005a,b; Šantrůčková et al., 2005). Estimates of the importance of soil CO2 fixation for the C balance in certain ecosystems or within an entire soil profile are rare (Ge et al., 2016; Yuan et al., 2012) and analyses of diversity and abundance of carboxylases are missing entirely. Soil OC becomes progressively enriched in 13C with increasing soil depth (Bird et al., 2002; Gentsch et al., 2015; Nadelhoffer and Fry, 1988; Torn et al., 2002). There are several explanations but no one can fully explain the measured isotopic shift. The enrichment of soil OC with depth can be connected with decrease of δ 13C of atmospheric CO2 by 1.3‰ due to Suess effect (McCarroll and Loader, 2004), with preferential decomposition of different organic compounds and microbial fractionation during litter decomposition or mixing of new C input with old soil OC (Buchmann et al., 1997; Ehleringer et al., 2000; Šantrůčková et al., 2000). Another hypothesis that has been discussed but never supported experimentally states that soil microbes should be isotopically heavier as a result of carboxylation reactions (Ehleringer et al., 2000). Whenever carboxylation reactions are involved, CO2 molecules used in the reactions likely originate from the soil atmosphere, which is isotopically heavier than the organic materials being decomposed (Cerling et al., 1991). The 13CO2 enrichment of bulk soil atmosphere is highest in the uppermost soil horizons, where CO2 originates mostly from the atmospheric air. In deeper horizons of the soil profile, CO2 originates from organic matter decomposition and carries the isotopic signal of decomposed material. But still CO2 remaining in the soil that surrounds microbes is 4.4‰ heavier than organic matter at the location due to slower diffusion of heavier 13CO2 than lighter 12CO2 (Cerling et al., 1991). CO2 hydrogenation causes further enrichment of 13C in HCO3 − by 8–12‰, depending on temperature (Mook et al., 1974).

On a global scale, the creation of artificial forest plantations on the initially forestless or deforested areas is regarded an important factor enhancing sequestration of carbon dioxide in different natural and climatic zones of the world [32, 35, 44, 50–52]. Estimates of the portion of carbon accumulated by the soil in forest plantations vary from less than 1% [47] to 39% [53]. Such a range of available estimates of the contribution of soils to the CO2 absorption is due to the strong dependence of the humus budget in forest ecosystems on a number of factors, including the initial humus content of the soil, the technology of its pre-planting preparation, the species composition and age of tree plantations, topography, climate, etc. [40, 43]. The most significant growth in the humus content of soils was observed during the forest planting or during natural overgrowing of abandoned arable lands with a low initial organic matter content [5, 38, 45, 46, 49, 50, 53]. At the same time, the planting of forests on pastures and in prairies led to some decrease in the storage of organic carbon and nitrogen in soils [39, 48]. Soil moistening also affects the dynamics of humus in soils under the forest. According to research in various regions of the world, a positive balance of humus is usually observed in the areas with precipitation of up to 800–1000 mm, and a negative balance of humus is observed in the areas with precipitation of more than 1000 mm [37, 38, 42, 49, 50]. Hence, forest planting may have a positive effect on the humus budget only for degraded soils of former agricultural lands under conditions of optimal moistening and the presence of grass cover in forest plantations. In addition to the impact on the dynamics of organic matter, forest planting significantly affects other soil indicators, including, in particular, the composition of the soil adsorption complex. A number of researchers note that an increase in acidity and a decrease in the amount of exchangeable bases are more expressed under coniferous species [33, 36, 39, 45, 46, 53]. The acidifying effect of forest plantations increases with an increase in the age of forest plantations and in the degree of their moistening [40]. Beginning from the early works by Korzhinsky [16], Tkachenko [29, 30], and Gorshenin [10], studies of the impact of artificial forest plantations on soil in the chernozemic zone have traditionally been focused on identification of the features of soil degradation. Numerous studies showed that the creation of forest plantations in the steppe and forest-steppe zones on plains with the hydrothermal coefficient of Selyaninov of 0.6–1.3 leads to certain changes in the properties of chernozems and dark chestnut soils. In particular, under the impact of woody vegetation, the crumb– granular structure is transformed into the angular blocky structure, the acidity of the soils increases, the humus content decreases, and the Cha/Cfa becomes smaller. Most scientists did not consider these changes as soil degradation [2, 3, 7, 20–22, 24] and emphasized the positive effect of forest vegetation on the original soil properties. The latter is manifested by an increase in the thickness of the humus horizon, an improvement of the soil physical and water-physical parameters, and leaching of soluble salts [3, 11, 13, 22, 23, 28, 31]. It was suggested that such soils should be referred to as forest-reclaimed chernozems [27].

Soil quality is one of the three components of environmental quality, besides water and air quality (Andrews et al., 2002). Water and air quality are defined mainly by their degree of pollution that impacts directly on human and animal consumption and health, or on natural ecosystems (Carter et al., 1997; Davidson, 2000). In contrast, soil quality is not limited to the degree of soil pollution, but is commonly defined much more broadly as “the capacity of a soil to function within ecosystem and land-use boundaries to sustain biological productivity, maintain environmental quality, and promote plant and animal health” (Doran and Parkin, 1994, 1996). As Doran and Parkin (1994) state explicitly, animal health includes human health. This definition reflects the complexity and site-specificity of the belowground part of terrestrial ecosystems as well as the many linkages between soil functions and soil-based ecosystem services. Indeed, soil quality is more complex than the quality of air and water, not only because soil constitutes solid, liquid and gaseous phases, but also because soils can be used for a larger variety of purposes (Nortcliff, 2002). This multi-functionality of soils is also addressed when soil quality is defined from an environmental perspective as “the capacity of the soil to promote the growth of plants, protect watersheds by regulating the infiltration and partitioning of precipitation, and prevent water and air pollution by buffering potential pollutants such as agricultural chemicals, organic wastes, and industrial chemicals” (National Research Council, 1993 as cited in Sims et al. (1997)). Soil quality can be assessed both for agro-ecosystems where the main, though not exclusive ecosystem service is productivity, and for natural ecosystems where major aims are maintenance of environmental quality and biodiversity conservation. Given the scope and readership of this journal, the “nonecological functions” of soil sensu Blum (2005), such as the physical basis of human activities, source of raw materials, and geogenic and cultural heritage, are beyond the scope of this review. Extrinsic factors such as parent material, climate, topography and hydrology may influence potential values of soil properties to such a degree (Fig. 1) that it is impossible to establish universal target values, at least not in absolute terms. Soil quality assessment thus needs to include baseline or reference values in order to enable identification of management effects. Soils often react slowly to changes in land use and management, and for that reason it can be more difficult to detect changes in soil quality before non-reversible damage has occurred than for the quality of water and air (Nortcliff, 2002). Therefore, an important component of soil quality assessment is the identification of a set of sensitive soil attributes that reflect the capacity of a soil to function and can be used as indicators of soil quality. Because management usually has only limited short-term effects on inherent properties such as texture and mineralogy, other indicators, including biological ones, are needed. The distinction between inherent (static) and manageable (dynamic) attributes, however, is not absolute and also context-dependent (Schwilch et al., 2016). For example, stoniness as an inherent property is nevertheless manageable, e.g. by removal of stones from an area to facilitate tillage and to build separating walls between fields, or by addition of gravel and stones to improve friability, to accelerate soil warming in spring or decrease evaporation. Soil management by humans has even given rise to separate classes in the soil taxonomic system, such as Plaggic anthrosols, the plaggen soils of northwestern Europe (e.g., Blume and Leinweber (2004)), and Terric anthrosols, the Amazonian Dark Earths, also known as Terra Preta de Índio (Glaser and Birk, 2012).