No significant transfer of the rare earth element samarium from spiked soil to alfalfa by Funneliformis mosseae

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REE transfer and toxicity to microorganisms

With more concern about the environmental impact of REE emission, more and more studies have been carried out on REEs and organisms. As an important biological component in the soil, microorganisms represent the largest biodiversity and play a crucial role in the biogeochemical cycle of the REEs. Previous studies showed that microorganisms can affect the mobility and the availability of REEs in the soil (Taunton et al. 2000; Perelomov and Yoshida 2008). However, very little information can be found related to REE effect, uptake and accumulation by microorganisms. Bergsten-Torralba et al. (2020) studied the toxicity of La, Nd and Sm on organisms including the fungi Penicillium simplicissimum and Aspergillus japonicus. Compared to the algae and crustaceans in the same experiment, the two fungi showed high resistance. Besides the studies on their resistance to REEs, microorganisms were also reported to affect the biogeochemical cycle of REEs. Fathollahzadeh et al. (2018) showed that Klebsiella aerogenes enhanced the bioleaching of REEs from monazite ore solutions. A few studies found that some bacteria (e.g. Pseudomonas aeruginosa, Variovorax paradoxus and Comamonas acidovorans) (Texier et al. 1999) and fungi (e.g. Amanita flavorubescens and A. rubescens, Russula pectinatoides,) (Aruguete et al. 1998; Kamijo et al. 1998) can absorb REEs in their cells and excrete REE- containing materials up to 10 mg/g. Yeast, as a model organism to study the metal toxicity and accumulation mechanisms, has been used in many REE studies. Different yeast strains were reported to have high ability to accumulate Nd (Palmieri et al. 2000; Vlachou et al. 2009). High doses of REEs can be toxic to microorganisms, still, a large number of microorganisms can adapt to the REE contaminated environment and even accumulate REEs.

REE transfer and toxicity to aquatic environment

The aquatic environment was considered a sink for contaminations including REEs (Gwenzi et al. 2018). The behaviour of REEs differs from element to element due to their physico-chemical properties. Gd based agents were reported having high stabilities and mainly exist in the dissolved form (Holzbecher et al. 2005; Knappe et al. 2005). On the contrary, anthropogenic La and Sm predominantly exist in the colloidal or nanoparticulate form according to the complexing agents and pH of the environment (Kulaksiz and Bau 2011).
In order to study REE toxicity different aquatic organisms have been tested. Using 3 REEs spiked in the aquatic system with algae, microcrustaceans and fungi, results showed that algae are very sensitive to REEs and can be used as a model organism to indicate the toxicity of REEs in the aquatic environment (Bergsten-Torralba et al. 2020). Spiking experiments were used to study the toxicity of different forms of REEs to algae, and showed that the water soluble REEs are more toxic than the acid extractable and the organic bound species (Hao et al. 1998). Besides algae, Daphnia, have also been used in different studies as a model organism to understand the toxicity and the biogeochemical cycle of REEs in the aquatic environment (Barry and Meehan 2000; Ma et al. 2016; Blaise et al. 2018; Galdiero et al. 2019). There is less information about REE toxicology to fish. Some REE risk assessment and bioaccumulation studies have been carried out on fresh water fishes (Yang et al. 2016a; Nørregaard et al. 2019). When it comes to marine fish species, studies mainly concern the analysis of REEs in fish tissues (Picard et al. 2002; Mayfield and Fairbrother 2015).

REE transfer and toxicity to plants

Even though REEs are not nutrient elements to plants, and not considered as essential elements for plants, positive effects of REEs at low concentrations were reported on plant growth (Diatloff et al. 1995c; Xu et al. 2002; Hong et al. 2003; Xiangsheng et al. 2006; Cheng et al. 2015). Since the early 1990s, REEs have even been used to promote plant growth and have been widely applied in agriculture in China. It has been reported that the yield increases from 5% to 15% due to REE addition for native crop species under varying soil and nutrient conditions (Hu et al. 2004).
However, when a high dose of REE fertiliser has been applied, plant growth was limited and even a toxic effect was found (Hong et al. 2003). Many studies have been carried out on the toxicity of REEs to plants. The concentration of REEs in plants growing in natural soil is rather low, ranging from 0.01 to 2 mg·kg-1 depending on the plants and the soils (Tyler 2004a). Despite the low REE concentrations in unpolluted soils, the phytotoxicity can become a concern in the mining, industrial or other contaminated areas (Carpenter et al. 2015). Spiking experiments have been performed to study the REE acute toxicity to plants. The early REE spiking studies have been carried out by Diatloff et al. (1995a, b) on La and Ce toxicity to corn and mungbean seedlings growing in nutrient solutions. The results showed a strong inhibition of mungbean root elongation and plant growth with Ce (even at micromolar level) while no significance with La. Similar toxicity study has been carried out in artificial soils with different pHs by Thomas et al. (2014). They used La, Ce and Y and all 3 elements showed inhibition effects to the germination and the growth of native and crop plants. The effects of REEs on 6 plant species have been studied by Carpenter et al. (2015). Seed germination rate was not affected by REE polluted soil; however, high levels of growth inhibition were found in the REE polluted soil, especially of native species. All species showed the ability to take up and accumulate REEs from soils in roots and shoots. Comparatively much higher tolerance to REEs was found in crop species than native species (Carpenter et al. 2015). Yuan et al. (2017) also studied the uptake and fractionation of REEs in plants and the influence of REEs on plant biomass. The results showed that the biomass and accumulation were enhanced at low concentration of REEs but inhibited at higher concentration.
Plants growing in polluted soils can absorb heavy metals and accumulate the metals in their cells. The term “hyperaccumulator” was used to describe plants which can grow on metalliferous soils and accumulate extremely high amounts of heavy metals in the aerial organs, far over the levels found in the majority of species, without suffering phytotoxic effects (Rascio and Navari-izzo 2011). Many plants have been studied and found having hyperaccumulating abilities. Pteris cretica var. nervosa (Zhang et al. 2017) and Pteris vittata (Liu et al. 2018; Evandro et al. 2019; Abou-shanab et al. 2020) were reported to accumulate arsenic (As). Many plants were reported to be Cd hyperaccumulators, such as Solanum nigrum L. (Sun et al. 2008; Dou et al. 2020), Lantana camara L.(Liu et al. 2019), Phytolacca americana (Zhao et al. 2011), Sedum alfredii (Lu et al. 2010). With the intense REE mining activity and environmental issue caused by REE wastes, more and more studies have been carried on REEs hyperaccumulators. The earliest REE hyperaccumulator has been reported by Robinson et al. (1938) and Robinson (1943) who showed that hickory leaves can accumulate REEs up to 2296 mg·kg-1 dry weight. Till now, many plants have been reported to be REE hyperaccumulator. Phytolacca americana was repeatedly reported to be an excellent REE hyperaccumulator (Yuan et al. 2017; Grosjean et al. 2019) with a high biomass and wide distribution all over the world (van der Ent et al. 2018), which can accumulate up to 4000 mg·kg-1 REEs in the dry weight (Yuan et al. 2017). Dicranopteris dicthotoma collected on-site also contained high concentrations of REEs (675-3358 mg·kg-1) in the dry leaves (Wang et al. 1997). Other than these plants, Carya tomentosa, Pronephrium triphyllum were also reported to be good hyperaccumulators with up to 1350 and 1027 mg·kg-1 REEs in their dry weight (Thomas 1975; Lai et al. 2005).
Beyond the inhibition of plant growth, REEs were also reported to influence the ionic fluxes into cells in different ways which may affect several plant processes. Previous studies on the effects of REEs on nutrient (N, P, and K) uptake by plants reported different conclusions (Hu et al. 2002a; Wang and Liang 2014; Chang et al. 2018; Grosjean et al. 2019). However, different analytical methods were applied by researchers and complex phenomena are involved in the influence of REEs on nutrient uptake (Hu et al. 2004).
The excessive amounts of REEs can be toxic not only to plants but also enter the food chain and accumulate in herbivorous consumers, higher class predators and even human beings (Loell et al. 2011a; Turra et al. 2013). Studying the transfer of REEs in the terrestrial environment, their effect on plants, and their transfer from soil to plants are important to understand the biogeochemical cycle of REEs, the effect of REEs to organisms and human health and also soil remediation.

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AMF promotion of plant growth and nutrient uptake

Growth promotion and nutrient uptake increase have been reported in many studies on different plants using different AMF species (Smith and Read 2008). In the study of Affokpon et al. (2011), 30 AMF strains have been inoculated to soils to study the growth of tomato and carrot. Among the 30 strains, AMF strain A. scrobiculata BEN202 was reported promoted carrot yields by 300% and tomato yields by 26% compared to the non-AMF control. Same growth promotion has been found in barley (Hordeum vulgare), cucumber (Cucumis sativus), field pea (Pisum sativum), tomato (Solanum lycopersicum) and wheat (Triticum aestivum cv. Kulin) (Johansen et al. 1992, 1993; Graham and Abbott 2000; Nagy et al. 2009; Smith and Smith 2011b; Mbodj et al. 2018).
The promotion of host plant growth is strongly correlated with the nutrient supply contributed by AMF. In fact, AMF significantly improve water and nutrient uptake, especially P uptake and transport, which is a diffusion limited nutrient in the soil. Many pieces of evidence showed that the main place of the symbiotic activity where the exchanges of water, nutrient and photosynthate sugar occurred is the arbuscule/cortical cell interface of AMF (Dalpé and Monreal 2004; Fitter et al. 2011). Several studies focused on the nutrient supply that has been improved by AMF and therefore improved plant growth and fruit yield. Charron et al. (2001a, b) have carried out experiments to study AMF effects on onions and the nutrient transfer. Even though the two AMF (G. versiforme and G. intraradices) have different abilities on the P and N transfer, both of them showed significant higher P and N concentration in plant tissue. AMF were reported paving the way for nitrogen-fixing bacteria to form intracellular symbioses with plant cells and therefore promoted better growth (Parniske 2008). Other than P and N, AMF also increased significantly the uptake of other mineral nutrients such as K, Ca, Mg, Fe, Mn and Zn. AMF also promoted P and Zn uptake in leeks and therefore improved leek growth (Sorensen et al. 2003). Ortas (2012) pointed out that AMF enhanced the concentration of P and Zn in green pepper, especially in nutrient-deficient conditions. However, Koide and Mosse (2004) mentioned that a growth depression can be found due to AMF inoculation for non-host species or host species when phosphate availability is high, while in most of the natural soils, P availability is a limiting factor for plant growth. Antunes et al. (2012) applied an AMF strain isolated from a nutrient-deficient soil to plants growing in different nutrition level conditions. They found that plant growth was largely promoted in the nutrient depleted soil while the plants showed no growth promotion and even inhibition with fertilization treatments.

Soil sampling and characteristics

In total 6 soils have been used in this PhD project including two uncontaminated soils and four contaminated soils collected from REE mine tailings sites and the surrounded unmined sites in China (Soil parameters in Table 2.1 & 2.2). In addition, one uncontaminated OCDE/ISO artificial soil has been used in the Chapter 6.
The two uncontaminated soils were sampled (5-25 cm depth) in North-Eastern France: a forest soil (Maron, 48°38’45.75″N, 6°4’48.44″E) and an agricultural soil (Bouzule, 48°44’7.95″N, 6°19’9.64″E). The two soils have different pH values and the physicochemical parameters of the two sampled soils can be found in the Table 2.1. The collected soils were dried in a dark and ventilated room until the weights remained constant. Dried soil samples were ground and sieved to 2 mm. After homogenization, the two soils were packed and stored in a dry and dark room ready for experiments. To avoid the compaction of Bouzule soil, it was partially amended with sand (sand/soil = 30w/70w) which has been pre-sieved (2 mm), washed and dried to allow better seedling growth. The two uncontaminated soils have been spiked with REEs in the Chapter 3, 5 and 6 to understand the toxicity of REEs to plants and the transfer of REEs in the terrestrial environment.
The soil parameters of the two uncontaminated soil were analysed by Laboratoire d’Analyse des Sols d’Arras (Table 2.1).

Total REE and major elements analysis in soil and plant

Soil. To analyze the REEs and other elements in soil, 200 mg of soil sample was transferred into a cleaned Teflon tube designed for CEM Mars 5 Digestion Microwave System® and 10 mL of Aqua regia. Digestion was processed under the adapted programme (15 min digestion at 25°C and 10 min digestion at 170°C). After the digestion and cooling down procedure, the digested samples were filtered (DigiFILTER, SCP Science® 010-500-070) and the solutions were adjusted into 10 mL in centrifuge tubes before the analysis.
Plant. Dry vegetal samples were digested in nitric acid (HNO3) and hydrogen peroxide (H2O2) solution (6 mL of 69% HNO3 and 3 mL of 30 % H2O2) (CAS: 7722-84-1). 200 mg of plant sample was transferred into a DigiTUBEs followed by a DigiPREP MS 48 & TS (SCP Science® 010-500-205 & 010-500-275) hot plate digestion. After cooling down, the solution was filtered, adjusted to 10 ml and kept at 4 °C until the analysis. REEs, nutrient element P and other major elements (K, Ca, Na, Mg) were analysed in the Chapter 4,5 and 6. For the two pot experiments, the chemical analysis was performed using inductively coupled plasma-optical emission spectroscopy (ICP-OES) analytical technique using a Varian Inc.® (Part A) – 720/730-ES ICP-OES. For the compartmented pot experiments, inductively coupled plasma mass spectrometry (ICP-MS) technique was used, with the certified reference material BCR®-667 estuarine sediment (rare elements, Th, U) as a reference. The detection limit for Sm concentration in the plant samples analyzed by ICP-MS was 0.0001 mg·kg-1.

Table of contents :

Chapter 1. State of the art and objectives
1.1 Rare earth elements (REEs)
1.1.1 What are REEs?
1.1.2 REE applications
1.1.3 Source of REEs
1.1.4 REE reserves and major mining deposits
1.1.5 REE production
1.2 REEs in the environment
1.2.1 REEs in soil
1.2.2 REE transfer and toxicity to microorganisms
1.2.3 REE transfer and toxicity to aquatic environment
1.2.4 REE transfer and toxicity to plants
1.2.5 REE transfer and toxicity to humans
1.3 Arbuscular mycorrhizae
1.3.1 Introduction
1.3.2 Effects of AMF on plants
1.3.3 Effect of AMF on plant in metallic polluted soils
1.4 Objectives
2 Materials and methods
2.1 Soil, plant and mycorrhizal fungi
2.1.1 Soil
2.1.2 Plants
2.1.3 Mycorrhizal fungi
2.2 Experiment design
2.2.1 Seed germination
2.2.2 Pot experiments
2.2.3 Compartment experiments
2.2.4 Growth conditions and watering
2.2.5 Spore germination tests
2.3 Analysis
2.3.1 Bioavailability of REEs in soil
2.3.2 Total REE and major elements analysis in soil and plant
2.3.3 Root mycorrhizal colonization
2.3.4 Data analysis
Chapter 3. Bioavailability and transfer of elevated Sm concentration to alfalfa in spiked soils .. 55 Bioavailability and transfer of elevated Sm concentration to alfalfa in spiked soils
Chapter 4. No significant transfer of the rare earth element samarium from spiked soil to alfalfa by Funneliformis mosseae
Chapter 5. Transfer of La, Ce, Sm and Yb to alfalfa and ryegrass from spiked soil and contribution of mycorrhizal inoculation with Funneliformis mosseae
Chapter 6. Availability and toxicity to AM fungi of REE contaminated soils from ionic mine tailings in China
Chapter 7. General discussion, conclusion and perspectives
7.1 General discussion
7.1.1 What are the bioavailability of REEs in the soil and their effects on plants?
7.1.2 Effect of an AM fungus on plants growing in an REE contaminated soil and on the transfer of REEs to the plant?
7.1.3 Effects of an AM fungus on plants growing in a soil with multiple REE contamination and on the transfer of these REEs to plants?
7.1.4 What are the bioavailability of REEs in historically contaminated mining tail soils and their effects on AMF?
7.2 Perspectives
7.3 Conclusion
References 

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