Mechanisms Behind Ph Changes by Plant Roots and Shoots Caused by Elevated Concentration of Toxic Elements

Mechanisms behind pH changes by plant roots and shoots caused by elevated concentration of toxic elements

Muhammad Tariq Javed

Doctoral Thesis in Plant Physiology at Stockholm University, Sweden, 2011

Mechanisms behind pH changes by plant roots and shoots caused by elevated concentration of toxic elements

Muhammad Tariq Javed

Cover illustration: Microscopic micrograph of Elodea canadensis leaf. Photo: Muhammad Tariq Javed

ISBN 978-91-7447-413-8, pp 1-40

Printed in Sweden by Universitetsservice, US-AB, Stockholm 2011 Distributor: Department of Botany, Stockholm University

I dedicate this work to my parents for their endless prayers

Abstract

Toxic elements are present in polluted water from mines, industrial outlets, storm water etc. Wetland plants take up toxic elements and increase the pH of the medium. In this thesis was investigated how the shoots of submerged plants and roots of emergent plants affected the pH of the surrounding water in the presence of free toxic ions. The aim was to clarify the mechanisms by which these plants change the surrounding water pH in the presence of toxic ions.

The influence of Elodea canadensis shoots on the pH of the surrounding water was studied in the presence of cadmium (Cd) at low initial pH (4-5). The involvement of photosynthetic activity in the pH changes was investigated in the presence and absence of Cd. The cytosolic, vacuolar and apoplasmic pH changes as well as cytosolic Cd changes in E. canadensis were monitored. The influence of Eriophorum angustifolium roots on the pH of the surrounding water was investigated in the presence of a combination of Cd, copper, lead, zinc and arsenic at low initial pH (3.5). Eriophorum angustifolium root exudates were analyzed for organic acids.

Elodea canadensis shoots increased the pH of the surrounding water, an effect more pronounced with increasing Cd levels and/or increasing plant biomass and increased plant Cd uptake. The pH increase in the presence of free Cd ions was not due to photosynthesis or proton uptake across the plasmalemma or tonoplast. Cadmium was initially sequestered in the apoplasm of E. canadensis and caused its acidosis. Eriophorum angustifolium roots increased the surrounding water pH and this effect was enhanced in the presence of arsenic and metals. This pH increase was found to depend partly on the release of oxalic acid, formic acid and succinic acid by the plants.

In conclusion, E. canadensis shoots and E. angustifolium roots were found to increase the low initial pH of the surrounding water. The pH modulation by these species was enhanced by low levels of free toxic ions in the surrounding water.

Keywords: Elodea canadensis, Eriophorum angustifolium, fluorescence microscopy, mechanisms, metals and metalloid, pH changes, protoplasts, phytofiltration, phytostabilization, root exudates.

List of papers

The present thesis is based on the work contained in the following papers, which are referred to in the text by their Roman numerals;

I. Javed, M.T., Greger, M., 2011. Cadmium triggers Elodea canadensis to change the surrounding water pH and thereby Cd uptake. International Journal of Phytoremediation. 13: 95-106. II. Javed, M.T., Lindberg, S., Greger, M., Cadmium induces cellular pH changes in Elodea canadensis and causes external basification. (Submitted). III. Javed, M.T., Stoltz, E., Lindberg, S., Greger, M., pH changes and organic acids exudation by Eriophorum angustifolium roots exposed to elevated concentration of toxic elements. (Submitted). IV. Javed, M.T., Lindberg, S., Greger, M., Cytosolic uptake of cadmium causes an extra- and intra-cellular basification in Elodea canadensis. (Manuscript).

My contributions to the papers were as follows; I was responsible for the writing of all papers with help from the co-authors. I planned the experiments in all papers with help and advice from the co-authors. I performed the laboratory work in all the papers (I was assisted with laboratory work and collection of Eriophorum angustifolium seeds). Reprint of Paper I was made with permission from the publisher.

Table of contents

Abstract ...... 7 List of papers ...... 8 Table of contents ...... 9 Abbreviations ...... 10 Introduction ...... 11 Phytoremediation and metal stress in plants ...... 11 Plant-induced changes in rhizosphere pH ...... 13 pH changes in water by roots and shoots of wetland plants after metal exposure ...... 15 Role of plant-induced pH changes in phytoremediation ...... 15 Objective ...... 18 Comments on methodology ...... 19 Choice of plants ...... 19 Cadmium concentration and initial pH used during the experiments ...... 19 pH measurements ...... 20 Loading of apoplasm and protoplasts with ion-specific dyes ...... 20 Metal and metalloid analysis ...... 21 Method development for isolation of E. canadensis protoplasts ...... 21 Results and discussion ...... 23 Cadmium and pH effect on the growth of E. canadensis ...... 23 Metal uptake in wetland plants ...... 25 Metal uptake response of E. canadensis and E. angustifolium exposed to metals……. and arsenic ...... 25 Effect of plant biomass on cadmium uptake in E. canadensis ...... 26 Effect of external pH on element uptake in E. canadensis and E. angustifolium ...... 26 Metal-induced pH changes in wetland plants ...... 27 External basification by E. canadensis shoots after cadmium exposure ...... 27 Photosynthesis and cadmium-induced external basification by E. canadensis ...... 28 Cytosolic proton dynamics in E. canadensis after cadmium exposure ...... 28 Apoplasmic cadmium-detoxification in E. canadensis and external basification ...... 29 Symplasmic cadmium-detoxification in E. canadensis and external basification ...... 30 Root-induced basification of the growth medium by E. angustifolium after metals….. and arsenic exposure ...... 32 Concluding remarks ...... 34 Future prospects ...... 34 Acknowledgements ...... 35 References ...... 36

Abbreviations

AMD Acid mine drainage AOE Antioxidative enzymes BCECF-AM 2′, 7′-Bis-(2-carboxyethyl)-5-(and-6)-carboxyfluorescein) acetoxymethyl ester CEC Cation exchange capacity 6-CFDA 6-Carboxyfluorescein diacetate DCMU 3-(3, 4-Dichlorophenyl)-1, 1-dimethylurea DMSO Dimethyl sulfoxide DW Dry weight EC50 Median effective concentration ER Endoplasmic reticulum FA Fatty acids FURA 2-AM Fura-2-acetoxymethyl ester FW Fresh weight MDA Malondialdehyde MES 2-(N-morpholino)ethanesulfonic acid MW Molecular weight NADP-ME Nicotinamide adenine dinucleotide phosphate-malic enzyme OAA Oxaloacetate OAC Osmotically active compounds PEG Polyethylene glycol PEP Phosphoenol-pyruvate PEPC Phosphoenol-pyruvate carboxylase PS Photosynthesis TBA Thiobarbituric acid TCA Tricarboxylic acid cycle TRIS Trishydroxymethylaminomethane

Introduction

Heavy metal pollution poses a great threat to the environment and human health worldwide due to the persistent nature and toxicity of heavy metals and their accumulation in the food chain (Bahadir et al., 2007). Both essential and non-essential metals are present in fresh water at low concentrations (Le Faucheur et al., 2006). Mallick and Mohn (2003) reported elevated concentrations of heavy metals in most parts of the world as a result of human activities. In most cases, heavy metals are discharged into the environment from various industries including smelters, electroplating cells, metal refineries, textile-, mining-, ceramic- and glass factories (Bahadir et al., 2007). Ca.1150 million tons of heavy metals (copper, lead, cobalt, zinc, cadmium and chromium) have been mined by man since the Stone Age (Sheoran and Sheoran, 2006). Mining activities disturb the land, surface water and ground water (Kalin, 2004). The waste of mining activities is called mine tailing and is usually stored in large impoundments. Mine tailings rich in sulphide minerals form acid mine drainage (AMD) through reaction with atmospheric oxygen and water (Rimstidt and Vaughan, 2003).

Acid mine drainage is characterized by low pH, high concentration of metals and sulphates due to oxidation of metal sulphides (Tsukamoto et al., 2004). Younger et al. (2002) reported its damaging impact when it is introduced in the ground water system. Acid mine drainage can be neutralized by use of chemicals such as lime, calcium carbonate, hydrated lime, caustic soda and soda ash, which result in the production of voluminous sludge containing ca. 5 % solids. The disposal of this sludge represents a further environmental problem and leads to additional costs (Fiset et al., 2003).

During the past few decades, research efforts have been focused on using wetlands, commonly known as biological filters, and associated processes to remove heavy metals from AMD as a cheap solution (Matagi et al., 1998). Wetlands have been used to purify AMD with few successes where metal uptake by plants usually accounted for a minor proportion of the total mass removed (Sobolewsk, 1999). Metal accumulation, metal accumulation patterns and metal translocation in aquatic plants have been the subject of many studies (Deng et al., 2004; Matthews et al., 2005). Besides direct uptake, wetland plants affect the efficiency of removal by increasing the residence time (Skousen et al., 1994), reducing water velocity, and ultimately enhancing sedimentation of heavy metals. Wetland plants have also been shown to decrease the redox potential of the growth substrate and prevent pH decrease in the sediment, ultimately binding elements to the substrate (Nyquist and Greger, 2009a). Submerged plants accumulate high amounts of metals (copper, zinc and lead) both from water and sediment (Fritioff and Greger, 2003) probably due to the fact that submerged plant species have a higher percentage of biomass in direct contact with polluted waters than emergent ones. Survival of submerged plants, however, was threatened under harsh wetland conditions due to iron precipitation on their surfaces (Nyquist and Greger, 2009a). Aquatic macrophytes can tolerate high metal concentration in their immediate surroundings (Otte et al., 2004). However, the pH response mechanisms to metal exposure are scarcely documented in wetland plants.

Phytoremediation and metal stress in plants

Since plants take up metals from their surroundings, those with a high uptake capacity are suitable candidates to be used to clean polluted waters. “The use of green plants and their associated micro-organisms, soil amendments and agronomic techniques to remove, contain, 11

or render environmental contaminants harmless” is termed phytoremediation (Cunningham and Berti, 1993). Phytofiltration is one phytoremediation method where submerged plant species accumulate toxic compounds from polluted water (Dushenkov et al., 1995). Phytostabilization is another phytoremediation method where metal tolerant plants are used to reduce the mobility of metals, thus leading to the stabilization of the metals in the substrate (Salt et al., 1995).

Plants take up metals via their roots and leaves depending on their growth media. In aquatic plants, metal uptake by roots is the main way for metal accumulation, except for fully submerged species. In case of roots, metals are first taken up into the apoplasm, and then transported further by the symplasmic or apoplasmic pathway. The movement of metals from the external solution into the apoplasm is a non-metabolic, passive process (Marschner, 1995). At the cellular level, the metal uptake mechanism is probably the same for leaf and root cells (Greger, 1999). Metals taken up into the apoplasm reach the cytoplasm through the plasma membrane mostly in the form of cations. Heavy metal transport at cell membrane can be mediated by substrate-specific transporters, i.e. HMAs (heavy metal-ATPases), NRAMPs (natural resistance-associated macrophage proteins), CDFs (members of the cation diffusion facilitator family), ZIPs (members of the Zrt-, Irt-like protein family) and cation antiporters (Hall and Williams, 2003). Cadmium is taken up via channels permeable for calcium and potassium (Lindberg et al., 2004). Inside the cytosol, some metals bind to the negative charges of various macromolecules, some soluble molecules, or to sulphur-rich polypeptides which participate in the transport of metals into the vacuole across the tonoplast.

Heavy metals have several toxic effects on plants, particularly when they are transferred to the aerial tissues. Although toxic effects can be measured at the cellular, as well as the biochemical level, most of the studies have focused on growth effects in response to contaminants in wetland plants (MacFarlane and Burchett, 2002; Peng et al., 2010).

The most pronounced effect of heavy metals at the cellular level is the alteration of cell membrane permeability for nutrients and the leakage of ions and solutes (Ros et al., 1990; De Vos et al., 1991). Heavy metals may induce changes to the plasma membrane. Two mechanisms are generally proposed; i. A direct effect on sulphydryl groups of membrane constituents by binding of heavy metal ions to SH groups (De Vos et al., 1991). ii. Direct or indirect free radical-mediated lipid peroxidation (Shi et al., 1993). Heavy metals are known to have a strong affinity for sulphydryl and carboxyl groups. In higher plants, the H+-ATPase of the plasma membrane is highly affected by heavy metals. In maize roots zinc, cadmium, mercury, copper and lead strongly inhibit the Mg+-ATPase activity and also affects ATPase-related processes, such as H+-ion efflux and trans-membrane and trans-root potentials as well as potassium uptake (Lindberg and Wingstrand, 1985; Kennedy and Gonsalves, 1987).

Padinha et al. (2000) measured the thiolic protein concentration, adenylate energy charge (AEC) and photosynthetic efficiency in plants grown under varying degrees of metal contamination. In Spartima maritima L., with increasing heavy metal concentrations, the leaf AEC ratio as well as photosynthetic efficiencies decreased, but the concentrations of thiolic proteins increased. These thiolic proteins, after binding the metals, transported them out of the cells and thus maintained low heavy metal concentrations in the leaf symplasm. This suggests that these proteins have a role in metal elimination mechanisms. Singh et al. (2008) found that when grown under higher cadmium concentrations, black gram (Vigna mungo L.) accumulated TBARS (Thiobarbituric Acid Reactive Substances) and H2O2 with decreasing

chlorophyll contents, stomatal conductance and carbonic anhydrase activity. All these factors ultimately reduced the photosynthetic activity in black gram. Plant net photosynthesis was inhibited by cadmium in tomato, probably due to structural changes in the chloroplast (Baszýnski et al., 1980). In Beta vulgaris, cadmium inhibited the net photosynthesis due to impaired electron transport and reduced Calvin cycle activity (Greger and Ögren, 1991).

Heavy metals inhibit the activity of many enzymes. The main mechanisms of enzyme inhibition by metals include high affinity of metals for sulphydryl groups and the displacement of metals ions from the enzyme structure (Van Assche and Clijsters, 1990). Displacement of Ca2+ by Cd2+ in the protein calmodulin led to an inhibition of the calmodulin- dependent phosphodiesterase activity in radish (Rivetta et al., 1997).

Plants have developed different mechanisms to overcome the stress imposed by high metal concentrations. Metal stress avoidance and metal tolerance mechanisms are sketched in Fig. 1.

Figure 1. Metal stress avoidance strategies in higher plants (left hand side) and metal tolerance mechanisms as highlighted by arrows (right hand side) leading to enhanced metal detoxification and repair mechanisms in plants (modified after Siedlecka et al., 2001).

Plant-induced changes in rhizosphere pH

Uptake of nutrients and metal cations by plant roots is responsible for substantial changes of the rhizosphere pH.

According to Song et al. (2004), the rhizosphere pH is formed from;

i. Imbalance between the absorption of cations and anions. ii. Generation of CO2 by rhizosphere respiration. iii. The secretion of organic acids, H+, and other chemical components from the roots.

Aspalathus linearis is a N2-fixing legume used for tea production, and grows in highly acidic soils (pH 3-5.3). Field and glasshouse studies revealed a significantly higher pH in the rhizosphere of this plant than in non-rhizosphere soils. The analysis of the effect of A. linearis cultivation on the pH of the nutrient solution revealed a marked increase of 2.8 pH units in the nutrient solution surrounding the roots of inoculated (nodulated) plants, compared to 1.5 pH - - units for un-inoculated control plants. Titrimetric methods showed that OH and HCO3 were the components of alkalinity in the nutrient solution surrounding roots of A. linearis, and were directly responsible for the increase in rhizosphere pH (Muofhe and Dakora, 2000).

Generally, the rhizosphere pH changes may depend upon the excretion of buffering agents - 2- (OH , CO3 ) by the plant roots. The source of nitrogen taken up by plants can stimulate the release of buffering agents by imposing a cation/anion imbalance. Villegas and Fortin (2001) + reported that NH4 -uptake decreases rhizosphere pH as protons will be used for the exchange, - - 2- while NO3 -uptake increases rhizosphere pH as the plants exchange nitrate with OH or CO3 (Nye, 1981) and nitrate is taken up in a proton-symport mechanism. However, the pH increase - + due to NO3 -uptake was lower than the pH decrease due to NH4 -uptake, i.e. soil pH value + near the root decreased by 1.1 pH units when NH4 was applied and increased by only 0.25 - units with NO3 application.

Moreover, organic acids released by the roots can also change the rhizosphere pH, but different studies suggested contradictory effects. Some studies suggested rhizosphere acidification by exudation of organic acids (Marschner, 1995) while other authors suggested that the rhizosphere pH could rise due to the organic anions (Jones and Darrah, 1994). The cytosolic pH is maintained at values close to 7.0, and at this condition the organic acids should be present as anions. Therefore, the organic acids could take up protons when entering the more acidic rhizosphere, leading to an increase in pH. However, this can only work if the rhizosphere pH is below 7.0, and the organic acids should not be exported in an H+-antiport mechanism (Fig. 2) (Jones and Darrah, 1994; Jones, 1998).

Figure 2. Release of different substances from plant roots that may affect soil pH (Nye, 1981; Jones and Darrah, 1994; Marschner, 1995; Villegas and Fortin, 2001). Uptake and release of different anions and cations depends on specific transporters.

pH changes in water by roots and shoots of wetland plants after metal exposure

Wetland plants have the ability to influence the surrounding water pH by their roots and shoots, which may help them to overcome heavy metal stress. The shoots of a submerged macrophyte Elodea canadensis increased the pH of the surrounding water, particularly in the presence of cadmium (Nyquist and Greger, 2009b). During photosynthesis, the leaves of submerged plants absorb CO2 from wetland waters, which increases the water pH, but during the night the pH will decrease again due to release of respiratory CO2 by the plant (Cronk and Fennessy, 2001). However, small pH changes caused by CO2 uptake or release are buffered by the carbonate/carbonic acid system. Zeng et al. (2008) observed that rice roots significantly increased the rhizosphere pH under chromium stress, particularly when the chromium level in the nutrient solution was 50 µM or above. Stoltz and Greger (2002) reported the ability of two wetland species, Eriophorum angustifolium and Eriophorum scheuchzeri, to maintain a high pH in the rhizosphere, even on AMD (Fig. 3). The pH rise after the uptake of toxic ions may - - 2- be due to cation/anion imbalance, release of OH or HCO3 or CO3 , or release of organic anions (as described in the previous section). The heavy metal stress inhibits the activity of H+-ATPase (Astolfi et al., 2003) and thus increases the apoplasmic pH. For example, cadmium stress inhibits the H+-ATPase activity of the plasma membrane of sugar beet roots (Lindberg and Wingstrand, 1985). Such a decrease in ATPase activity should result in decreased proton efflux, which ultimately would cause an increase in rhizosphere pH.

Figure 3. Long-term effects by plants on the pH of the drainage water of Kristineberg tailings with low buffering capacity. SS=sewage sludge, E.a. = Eriophorum angustifolium, E.s. = Eriophorum scheuchzeri (Stoltz and Greger, 2002).

Role of plant-induced pH changes in phytoremediation

The pH of the surrounding medium is one of the important factors affecting the availability of metals to plants in both terrestrial and aquatic systems. But this effect can be different for terrestrial and aquatic plant species depending on the surrounding medium (water or soil) and the type of metal.

In hydroponics, metal uptake usually increases with increasing pH (Marschner, 1995) because at high pH fewer protons will compete with metal ions at uptake sites. Nyquist and Greger (2009b) reported that pH increase of the surrounding medium increased the cadmium uptake by shoots of E. canadensis and Carex rostrata probably due to diminished interactions between H+ and Cd2+.

Youssef and Chino (1989) obtained a depletion of copper over 2-3 mm in the rhizosphere solution due to basification of the rhizosphere induced by barley roots. These authors were not able to conclude if root-induced basification depended on metal uptake, and ultimately did not determine which processes controlled the metal depletion in the rhizosphere. Bravin et al. (2009) quantified how root-induced basification controlled the establishment of copper gradients in the rhizosphere of wheat grown in highly acidic, copper-contaminated soil. These authors observed that the total copper concentration decreased 3-fold in the rhizosphere solution due to root-induced basification.

Dispersal of mining waste in the surrounding areas by percolation, drainage or accidental breakdown of dams, is highly unwanted. Although a low pH only has temporary effects (as natural edaphic and biological processes will increase the pH again), the residual effects of metals may be long term, preventing the establishment of vegetation in the area. Plants have the ability to stabilize toxic metals in the tailing reservoir, and this ability may depend on plant root-induced pH-rise. Stoltz and Greger (2002) studied the influence of two cotton grass species on metal dispersal from limed and unlimed submersed mine tailings. These authors concluded that plants stabilize cadmium, copper, lead and zinc of the drainage water in unlimed mine tailing by maintaining a pH of ca. 6, and that the pH was decreased in limed mine tailing.

Heavy metal uptake from the soil increases with decreasing soil pH, because low pH increases metal availability. Protons have a higher affinity for negative charges on the colloids, thus they compete with metal ions of these sites and release metals. The soluble metal forms of both cadmium and zinc were greatly increased with decreasing soil pH, which significantly enhanced metal uptake in Thlaspi caerulescens (Wang et al., 2006). But this does not always happen. Metal uptake by plant roots at low pH can be low because of interaction between metal ions and protons (Fig. 4). Eriksson (1989) reported decreased cadmium contents in ryegrass when the soil pH was changed from 5 to 4 by adding sulphur. This can either be due to the interaction effects between Cd2+ and H+ at the root surface or root damage at low pH.

+ H Root

Soil colloid Me+ - - - - -

- - - - -

Figure 4. The release- of -metal - - - - -ions from soil colloid by protons at low pH and their uptake by plant roots (modified after Greger, 1999). Metal uptake can be high due to more available metal or low due to an interaction effect between metal and hydrogen ions.

The plant-induced pH rise is important for metal uptake from polluted waters. Such pH changes by plants can be of practical importance for phytofiltration as well as phytostabilization when dealing with acidic, metal-contaminated waste. Plant roots can

decrease the soil/sediment pH by releasing protons, or by redox reactions (including sulphide and iron oxidation) resulting in greater metal bioavailability (Jackson et al., 1993). This feature can be of particular interest for plants grown for phytoremediation purposes in alkaline medium. pH-based remediation offers more advantages for the use of submerged plants due to their high metal uptake potential as compared with terrestrial plants (Greger, 1999). Therefore, the establishment of aquatic-plant communities in waters polluted with metals, and characterized by low pH, will optimize the remedial potential. Different plant species behaved differently when exposed to a range of different pH values. Therefore, to optimize the use of plant-based remediation, there is dire need to understand the mechanisms behind plant- induced pH-changes.

Objective

In order to use wetland plants for phytofiltration and phytostabilization, it is of great importance to understand the plant´s responses to stress caused by toxic elements. One such response observed in Elodea canadensis was the pH increase under cadmium stress. Another observation was the maintenance of a high pH by Eriophorum angustifolium on AMD. Therefore, the overall goal of these studies was to elucidate the mechanism behind pH-rise in the surrounding medium caused by shoots of E. canadensis and by roots of E. angustifolium in the presence of different toxic ions.

The specific aims were to investigate the following points:

i. How the toxic ions trigger E. canadensis and E. angustifolium to alter the surrounding water pH and how the plant-induced pH changes affect the uptake of the toxic ions by the plants. ii. Whether the pH-changing capability of E. canadensis is affected by plant biomass as well as by the metal concentration in the surrounding medium. iii. The role of photosynthetic activities in plant-induced pH changes by shoots of E. canadensis in the surrounding medium. iv. If the pH increase of the surrounding medium depends on the uptake of protons into the cytoplasm and/or vacuole of E. canadensis. v. The role of organic acids in the root exudate in root-induced pH changes by E. angustifolium after exposure to toxic metal ions.

An understanding of the plant-induced pH-rise and its mechanism in plants exposed to heavy metal stress may illustrate fundamental aspects of plant stress physiology, and present tools in optimizing strategies for handling of water pollution, as well as crop establishment under stress conditions.

Comments on methodology

Choice of plants

The plants used for the studies were Elodea canadensis and Eriophorum angustifolium. These specific plants were chosen because they have the ability to change the surrounding medium pH in the presence of free metal ions (Stoltz and Greger, 2002; Nyquist and Greger, 2009b). Furthermore, laboratory and field studies with these species have demonstrated their potential use in removing and stabilizing metals and metalloids in polluted environments in cold climate zones.

Cadmium concentration and initial pH used during the experiments

The cadmium concentration used in the experiments was 0.5 µM. This concentration was chosen from toxicity test based on chlorophyll contents (Fig. 5). To decide on the initial pH to be used in the experiments, the influence of different pH values (2, 3, 4, 5, 6, 7, 8) on plant fresh weight, chlorophyll contents, catalase activity, malondialdehyde (MDA) contents and ascorbate peroxidase activity was studied. Malondialdehyde (a product of lipid peroxidation) contents were determined by reaction with thiobarbituric acid (TBA). These parameters suggested that E. canadensis can function normally at pH values between 3.3 and 5.65 (Fig. 6; Fig. 8).

Figure 5. Effect of different cadmium concentrations [0 (Ο), 0.01(Δ) 0.05 (□), 0.1 (ӿ), 0.25 (×), 0.5 (◊), 0.75 (–), 1 (+), 2 (-) µM] on chlorophyll contents of E. canadensis at neutral pH. The solid line is the calculated Weibull curve. The experimental data used for the calculation of Weibull curve are shown by different symbols. n= 3.

A modified Weibull frequency distribution (Taylor et al., 1991) was used to model the toxic effects of cadmium and acidity because it is simple and easy to understand. Median effective concentration (EC50) was calculated as the concentration of cadmium and protons that results in a growth reduction of 50% (of the response parameter, i.e., FW change, chlorophyll contents, MDA contents, catalase and ascorbate peroxidase activities) by the following formula;

1/d EC50 = c[(-ln 0.5) ]

Figure 6. Shoots of E. canadensis subjected to growth media with different pH for 3 days.

pH measurements

In different experiments, pH values were double checked using different pH-meters to avoid the probability of error. A portable pH-meter Schott handylab multi 12 (Schott Instruments, Mainz, Germany) was used during one experiment and a Metrohm-744 pH meter (Metrohm AG, Herisau, Switzerland) was used during the other experiment. Cytosolic and vacuolar pH changes in E. canadensis were investigated by fluorescence microscopy using the pH dependent dyes, 2′, 7′-bis-(2-carboxyethyl)-5-(and-6)-carboxyfluorescein) acetoxymethyl ester (BCECF-AM) and 6-carboxyfluorescein diacetate (6-CFDA), respectively.

Loading of apoplasm and protoplasts with ion-specific dyes

To monitor the dynamics of cytosolic ions - protons, Ca2+ and Cd2+ - under cadmium stress, isolated protoplasts were loaded with the ion-specific dyes BCECF-AM (Sigma-St Louis, MO, USA) for H+, Fura-2-acetoxymethyl ester (Fura 2-AM, Molecular Probes, Leiden, Netherlands) for Ca2+, LeadmiumTM Green AM (Molecular Probes, Invitrogen, Carlsbad, CA, USA) for Cd2+. To measure the vacuolar proton concentration, the isolated protoplasts were loaded with 6-CFDA (Sigma). Apoplasmic H+ concentrations were measured by loading the excised leaves of E. canadensis and roots of E. angustifolium (Fig. 7) with 5 µM Oregon Green 488-dextran (MW=10,000; Invitrogen) after heavy metal treatment by a vacuum infiltration technique (Geilfus and Mühling, 2011).

Figure 7. Micrograph of a root section of E. angustifolium loaded with 5 µM Oregon green 488-dextran photographed under fluorescent light using a fluorescein isothiocyanate (FITC) filter.

Before loading any dye, the protoplasts were washed twice in the loading medium (Medium A) containing 0.7 M sorbitol (Sigma), 1 or 0.1 mM CaCl2, 0.2% (w/v) polyvinylpolypyrrolidone (PVP, Sigma) and a buffer (pH 5.5) containing 5 mM trishydroxymethylaminomethane (TRIS, Labassco, Molndal, Sweden) and 5 mM 2-(N- morpholino)ethanesulfonic acid (MES, Sigma).

The BCECF-AM dye solution was prepared by mixing 5 µL of stock solution (1.6 mM) with 5 µL of ethanol (99%) and 1.25 µL pluronic F-127 (Sigma). From the mixed dye solution of BCECF-AM, 5 µL was added to 1 mL of protoplast suspension to obtain a final concentration of 3.54 µM. The Fura 2-AM solution was prepared by mixing 2 µL of Fura 2-AM stock solution (5 mg/mL) in dry (<0.1% v/v water) DMSO (dimethyl sulfoxide), 1.25 µL of pluronic F-127 solution (20% w/v) in DMSO and 6.75 µL of ethanol (99.5% v/v) (Sebastiani et al., 1999). From the above describe Fura 2-AM dye solution, 5 µL were added to 1 mL of protoplast suspension. A stock solution of LeadmiumTM Green AM dye was prepared by adding 50 µL of DMSO to one vial of the dye. This stock solution was then diluted 1:10 with 0.85% NaCl. Five microliters of dye solution along with 1.25 µL pluronic F-127 were added to 1 mL of protoplast suspension. For vacuolar 6-CFDA loading the protoplasts were stained with 3 µL of 5 mM 6-CFDA in DMSO through polyethylene glycol (PEG)-mediated osmotic shock (Seidel et al., 2005) giving a final concentration of 1.88 µM of 6-CFDA in the protoplast suspension.

Dye loading was performed in darkness in medium A at room temperature for 2 h with BCECF-AM and LeadmiumTM Green AM and for 2 to 3 h with Fura 2-AM at 22 oC. After loading, the samples were centrifuged and pellets were re-suspended in 1 mL of a solution similar to medium A, but with TRIS-MES buffer of pH 7 instead of pH 5.5 (medium B). Before measurements, samples were kept in darkness at room temperature for 25 min.

Metal and metalloid analysis

The metal and metalloid contents of dried plant samples were analyzed with a SpectrAA-100 atomic absorption spectrophotometer (Varian, Springvale, Australia). Flame spectrometry was used for analysis of cadmium, copper, zinc and lead. However, some of the samples were analyzed using furnace technique (GTA-97). The hybrid vapor generation technique was used for arsenic analysis. To eliminate errors based on interaction with the sample matrix, standards were added to the samples.

Method development for isolation of E. canadensis protoplasts

Leaves of E. canadensis consist of only two layers of cells (Paper II) and exhibit a light- induced polar pH reaction (Elzenga and Prins, 1989). Compared to terrestrial plants, the isolation of protoplasts from leaves of submerged species is inherently difficult. In case of terrestrial plants, protoplasts can be prepared from the mesophyll after the epidermis has been removed by stripping. Since the leaves of submerged plants consist of only two cell layers, stripping the epidermis is not possible. Different mechanical ways are available for releasing protoplasts from plant tissues but they have to be combined with enzymatic breakdown of cell walls. Enzymatic protoplast isolation is beneficial because it is quick and the number of released protoplasts is high. The release of protoplasts from E. canadensis leaves required an extra enzyme to be added to Cellulase and Macerozyme, due to the extra sugar component in

its cell wall. Therefore, the extraction of E. canadensis protoplasts required a mixture of 1.5% Cellulase from Trichoderma reesei (Sigma), 1.5% Hemicellulase from Aspergillus niger (Sigma), 0.75% Macerozyme R-10 from Rhizopus (SERVA, Mannheim, Germany) and 2.5% (v/v) β-glucuronidase from Helix pomatia (Sigma). According to our knowledge, this recipe was used for the first time.

Results and discussion

Cadmium and pH effect on the growth of E. canadensis

In order to determine the optimum pH and cadmium exposure conditions at which Elodea canadensis can perform non-stressed growth, cadmium and proton toxicity was analyzed during growth at different cadmium and proton concentrations. The results of cadmium toxicity test showed that the chlorophyll contents of E. canadensis started to decrease above cadmium concentrations of 0.5 µM in the nutrient solution (Fig. 5).

Median effective concentration, the concentration of cadmium that results in a growth reduction of 50%, was 0.57 µM. Elodea canadensis can function normally at cadmium concentrations 0.01-0.04 µM (Sun et al., 2011). Cadmium concentration of 0.5 µM was used in experiments which were in the range of toxicity. The choice of initial pH (4.0) used in the experiments in paper 1 was not suitable due to its toxic effects and probably the reason for decreased plant fresh weights observed even in the absence of any added cadmium.

The toxic effects of cadmium on E. canadensis depended more on reduced growth than on reduced photosynthetic activity, as shown by the increased Dry Weight:Fresh Weight (DW:FW) ratio (Paper 1). The increase in DW:FW ratio was the result of a marked decrease in the FW of the plants. The decrease in FW might be the effect of cadmium on chlorophyll contents which are slightly decreased at cadmium concentrations of 0.5 µM (Fig. 5). The growth inhibition of E. canadensis by cadmium may arise from its detrimental effects on light induced proton pumps resulting in suppressed cell wall extensibility and cell elongation rates as shown for Impatiens balsamina stem cells (Aidid and Okamoto, 1993). After prolonged cadmium exposure E. canadensis plants showed shorter and poorly pigmented shoots with thinner stems and less expanded leaves (Vecchia et al., 2005). Cadmium-induced growth inhibition in plants has been found to be concentration-dependent (Siroka et al., 2004).

The pH could also have a negative influence on the plant growth and activity. Different parameters were affected differently by the proton concentration in the medium. Catalases activity, FW, chlorophyll contents, MDA contents and ascorbate peroxidases activity was reduced to 50% at proton concentrations that corresponded to a pH of 5.65, 3.65, 4.05, 3.27 and 3.76, respectively (Fig. 8). Among the parameters tested for different pH values, catalase activity was more sensitive to acidity. The studied parameters suggest that the pH value which causes 50% reduced growth of E. canadensis lies between 3.3 and 5.65. This indicates that the decrease in FW in the absence of cadmium (Paper I) was probably due to the low pH (pH 4.0). Low pH, however, neither enhance the activity of catalases and ascorbate peroxidases nor MDA contents in E. canadensis showing that it does not have the same effects as other stresses (Fig. 8).

Figure 8. Effect of different proton concentration on catalase activity (A), fresh weights (B) chlorophyll contents (C), malondialdehyde (MDA) contents (a product of lipid peroxidation) (D) and ascorbate peroxidase activity (E) of E. canadensis. The solid line is the calculated Weibull curve. The experimental data used for the calculation of Weibull curve are shown by different symbols. In A, B and C, proton concentrations were 10-8 (Ο), 10-7(Δ), 10-6(□), 10-5(ӿ), 10-4(×), 10-3(◊), 10-2(-) M. In D and E proton concentrations were 10-7(Ο), 10-6(Δ), 10-5(□), 10-4 (ӿ), 10-3 (-) M. n=3.

Metal uptake in wetland plants

Metal uptake response of E. canadensis and E. angustifolium exposed to metals and arsenic

A linear relationship was found between cadmium concentrations in the growth medium and cadmium concentrations of E. canadensis shoots (Paper I) which corroborates the findings of Maine et al. (2001). There are different ways by which plant respond to heavy metal exposure (Baker et al., 1990). Plants may secrete various exudates which chelate metals with the substrate and inhibit plant metal uptake. However, this strategy seems not to be applied by E. canadensis because when an efficient avoidance strategy is active, the tissue concentrations of metal would not increase with the external heavy metal concentration. Furthermore, it was shown that exposure to 0.5 µM cadmium for 72 h did not increase secretion of organic acids from this species (Nyquist and Greger, 2009b).

Increased shoot cation exchange capacity (CEC) after cadmium exposure indicates the development of new binding sites in E. canadensis (Fig. 10b) and can be plant response to prevent cadmium uptake into the cytoplasm. Enhanced shoot CEC may also account for increased cadmium uptake with increasing cadmium in the medium as the uptake of divalent cations increased with increasing CEC of cell walls (White and Broadley, 2003). A unique feature of E. canadensis cell walls is that it contains extra sugars (Paper II & Paper IV). The presence of extra polysaccharides might be connected with the accumulation of increased amounts of cadmium in the cell walls of E. canadensis. Earlier works have shown that cadmium exposure resulted in the appearance of phenolic compounds in this species (Vecchia et al., 2005) which may account for increased cadmium uptake by chelating it.

The effective uptake of cadmium decreased with increasing external cadmium concentration (Paper I), which was consistent with the findings of Greger (1999). This may be due to high competition of metal ions at the uptake sites, as metal concentration per absorption area will be high with increased metal concentrations in the solution.

The roots of Eriophorum angustifolium respond to metals and arsenic exposure by releasing different organic acids (Paper III). Concentrations of arsenic and lead in roots and shoots of E. angustifolium increased but plant cadmium and copper contents decreased with increasing element treatment (Paper III). This differential element uptake may depend upon the different uptake pathways for different elements or competition between different elements at uptake sites. The elevated concentration of toxic elements (arsenic, lead and cadmium), may also decrease the uptake of essential micronutrients, i.e. copper and zinc. The total concentrations were significantly higher in roots than in shoots for all metals and metalloid examined i.e. arsenic, lead, copper, cadmium and zinc (Paper III). This indicates that most of the elements absorbed from the nutrient solution would remain in the roots. The exclusion of toxic elements from aboveground tissues may be a strategy to tolerate toxic elements (Taylor and Crowder, 1983). This mechanism seems to exist in E. angustifolium and may be due to metal accumulation in the root cell walls or root vacuoles.

Since the translocation factor (TF) value is less than 1 for all treatments at different concentrations, it might be logical to suggest this species for phytostabilization of different elements (Fig. 9).

Figure 9. Element translocation to shoots of E. angustifolium after five days exposure to different elements, copper, lead, zinc, arsenic (0, 15, 25, 35, 50 µM), and cadmium (0, 1.5, 2.5, 3.5, 5.0 µM) in hydroponics. Different letters indicate significant difference (P ≤ 0.05). n = 5, means ± SE.

Effect of plant biomass on cadmium uptake in E. canadensis

With increasing plant biomass of E. canadensis, the shoot cadmium concentration, as well as the effective uptake, significantly decreased (Paper I). Cadmium uptake in terrestrial plants increases with increasing plant absorption area (Greger, 1999) but that was not the case for E. canadensis. One reason could be that uptake via shoots functions differently as compared with root uptake. Moreover, at higher biomass, the metal uptake efficiency seems to be diminished due to plant competition for uptake of elements (Marschner, 1995). It may also be due to biological dilution depending upon proportionally larger absorption area in relation to the external metal concentration. This in turns leads to low internal cadmium concentration, as also demonstrated for cadmium uptake in Pinus sylvestris (Ekvall and Greger, 2003).

Effect of external pH on element uptake in E. canadensis and E. angustifolium

Cadmium uptake increased in E. canadensis with increasing pH in the culture medium (Paper I). The increasing metal concentration in roots of E. angustifolium after metal exposure can also be the effect of root-induced basification in addition to increasing metal concentrations in the external medium (Paper III). Increased metal uptake at high pH is likely due to diminished interaction between protons and metal ions at plant uptake sites (Marschner, 1995). In addition, the pH of the medium may have a marked influence on metal speciation (Sauvé et al., 2000) and this could enhance metal uptake in plants at high pH in solution. However, in case of sediment, decrease in the rhizosphere pH increases the solubility of metals which may

enhance metal uptake in the plant. The medium pH influenced arsenic mobility and plant uptake differently as compared to other metals depending on its different chemical properties (arsenic is a metalloid and exist as anion; Sracek et al., 2004). In solutions with low pH, uptake of arsenic can be high as arsenite, which is more mobile, is the dominant form. In solutions with high pH, its uptake can be low as arsenate, which is less mobile than arsenite, is the dominant form.

Metal-induced pH changes in wetland plants

External basification by E. canadensis shoots after cadmium exposure

The shoots of E. canadensis increased the surrounding medium pH when the pH was initially low (pH 4.0) (Paper I). The pH increase response became more prominent with increasing cadmium chloride concentration in the medium and/or with increasing plant biomass (Paper 1). This pH response was also shown with other cadmium salts [CdSO4 and Cd(NO3)2)], yielding the maximum pH increase when cadmium was added as cadmium nitrate (Fig. 10a).

Figure 10. The pH change of the surrounding medium in the presence of E. canadensis shoots (a) and shoot cation exchange capacity (b) subjected to 0.5 µM cadmium by use of different cadmium salts [(Cd(NO3)2, CdSO4 and CdCl2)] at an initial pH of 5.0 for 3 d. Different letters indicate the significance of differences (P ≤ 0.05). n = 3, ± SE.

The large increase in external pH when cadmium was supplied as cadmium nitrate (Fig. 10a) - - - may be due to NO3 -uptake which is coupled with release of OH . Further, the protons/NO3 symport as shown for in maize roots (McClure et al., 1990; Tsay et al., 1993) can be the reason for large increase in external pH.

Since proton uptake by the plants increased with increasing cadmium concentration in the medium (Tab. 1), the free cadmium ions are likely to activate the E. canadensis plants to buffer the surrounding water pH. Proton uptake by E. canadensis decreases with increasing

biomass (Tab. 1), which is probably due to biological dilution. This suggests that in this species, enhanced medium basification with increasing cadmium concentration depends on proton uptake. However, the increased pH rise with increasing plant biomass does not depend on proton uptake but is probably due to cadmium detoxification mechanisms (discussed in next sections). Furthermore, the increasing proton and cadmium uptake with increasing cadmium concentration in the medium (Tab. 1) suggests a Cd2+/H+ and/or Cl-/H+ symport in this species. Chloride part will come from added cadmium chloride.

Table 1. Increase in external medium pH, proton and cadmium uptake by different shoot biomasses of E. canadensis after 3 d in the presence of different added cadmium chloride concentrations (0, 0.1 and 0.5 µM) (Paper I). ______No. of shoots Parameter Cadmium treatment (µM) ______0 0.1 0.5 ______pH increase (units) 1 0.08 0.26 0.54 3 0.26 0.56 0.73 6 0.54 0.44 1.46

Cadmium uptake (µg / g DW) 1 79 516 2427 3 44 140 819 6 58 109 422 Proton uptake (µmol / g FW) 1 76 96 117 3 58 75 92 6 43 42 51 ______

Photosynthesis and cadmium-induced external basification by E. canadensis

The results suggested that besides normal photosynthetic activity, E. canadensis have some additional mechanisms, which consume H+, or release OH-, and thereby increase the medium pH (Paper I & IV). Since a maximal pH increase as compared to control was observed in darkness and in the presence of cadmium either in darkness or under PS inhibitor (DCMU) treatment, these additional mechanisms were stimulated when the photosynthetic activity was inhibited either by cadmium, darkness or photosynthetic inhibitor (Paper IV). Cadmium itself is a highly toxic metal to the photosynthetic process and the pH increase was found to be more prominent in the presence of the photosynthetic inhibitor (DCMU) and cadmium. Therefore, it is likely that such pH-changes induced by E. canadensis are a typical stress response, at least during growth in an acidic environment.

Cytosolic proton dynamics in E. canadensis after cadmium exposure

In order to investigate whether the basification of the surrounding medium depends on proton uptake across plasmalemma or tonoplast, the cytosolic and vacuolar pH of the E. canadensis

were monitored after the addition of cadmium. Cadmium had been hypothesized to cause cytosolic acidosis as reported by earlier investigations (Lindberg et al., 2007; Nocito et al., 2008) depending upon impaired cell membrane permeability after cadmium exposure at low pH. The results indicated that cadmium was sensed inside the cytoplasm of E. canadensis, and subsequently induced cytosolic, as well as vacuolar basification (Paper II). This may depend on the special type of plants used during this study (submerged plants) as compared to other reported cases with terrestrial plants. After cadmium exposure at low external pH, the apoplasmic pH decreased and cadmium binding to the cell wall was increased (Paper II). Probably, cell walls loosening occur due to enhanced cell wall acidity and enhanced cadmium binding to cell walls. This characteristic could help the plant to sequester cadmium outside the cytosol, thus protecting inner cell parts. The cytosolic pH increase under cadmium stress can be a trait of heavy metal tolerant species; e.g. sodium stress has been reported to cause a cytosolic pH increase in the salt-tolerant rice cv. Pokkali (Kader et al., 2007) and in salt- tolerant quince (D’Onofrio and Lindberg, 2009).

Apoplasmic cadmium-detoxification in E. canadensis and external basification

Clarifying the rate of uptake and transport of cadmium within the plants is essential to better understand the physiology of cadmium accumulation. Specifically, only few studies consider the Cd2+/H+ dynamics in aquatic macrophytes. In E. canadensis cadmium uptake by shoots exhibits two phases, i.e. apoplasmic binding and after some lag time symplasmic uptake (Paper IV). This is in accordance with the findings of Zhao et al. (2002) who reported that root uptake of divalent cations typically exhibits two phases: apoplasmic binding and symplasmic uptake. Elodea canadensis shoots initially sequestered cadmium in their apoplasmic region (Paper IV). During uptake studies, enhanced cadmium binding to the cell walls of E. canadensis as compared with other plant parts is a permanent property of this species as demonstrated by short term (Nyquist and Greger, 2009b) and long term studies (Vecchia et al., 2005). In contrast, for the terrestrial herb Sedum alfredii it had been shown that the apoplasm of the non-hyperaccumulating ecotype (NHE) bound more cadmium than the hyperaccumulating ecotype (HE) of S. alfredii which accounted for initial enhanced cadmium uptake in roots of the NHE compared to the HE ecotype. However, long term cadmium accumulation depended on the enhanced root-to-shoot translocation efficiency in the HE ecotype (Lu et al., 2008). After cadmium exposure, the apoplasmic pH decreased in E. canadensis (Paper II). Cadmium was initially sequestered in the apoplasm of this species (Paper IV). It is likely that E. canadensis binds more protons, as well as cadmium ions, to its cell walls at the initial stages.

Cadmium and proton uptake follows a similar dynamic in E. canadensis, but proton uptake was higher than cadmium uptake at a specific cadmium concentration in the medium (Tab. 1). This mechanism probably causes the surrounding medium pH rise in the presence of cadmium, as maximum pH increase in the surrounding medium was observed during the first 48 h of cadmium exposure (Paper 1). Enhanced cadmium binding to the cell wall of E. canadensis can be explained by the presence of matrix polysaccharides, particularly by pectic substances in its cell walls as revealed by cyto-chemical studies (Paper IV). Trapping of cations by cell wall polysaccharides has been shown in algal species, Vaucheria and Rhizoclonium (Crist et al., 1994).

Symplasmic cadmium-detoxification in E. canadensis and external basification

Beside apoplasmic cadmium sequestration in E. canadensis, the involvement of other mechanisms cannot be excluded. The external basification was enhanced under conditions when the plant was stressed either by cadmium or by other photosynthetic stress conditions (Paper I & IV). Cadmium exposure as well as photosynthetic inhibitor (DCMU) resulted in a breakage of photosynthetic electron transport inhibiting the reduction of NADP+ to NADPH. Under these conditions, there is a possibility of enhanced expression of phosphoenol-pyruvate carboxylase (PEPC) and NADP-dependent malic enzyme (NADP-ME), which are important for maintaining cytosolic pH and the antioxidant machinery. Phosphoenol-pyruvate carboxylase catalyzes the carboxylation of phosphoenolpyruvate (PEP) to oxalo-acetate (OAA) and inorganic phosphate (Pi) in the presence of bicarbonates. Nicotinamide adenine dinucleotide phosphate-ME de-carboxylizes malate to pyruvate, CO2 and NADPH. Protonation of Pi and diffusion of CO2 to the Calvin cycle may contribute to cytosolic basification upon cadmium treatment after lag time (Fig. 11). Photosynthetic counterparts of PEPC and NADP-ME are component of C4 photosynthesis. Therefore, switching of E. canadensis from C3 to C4 cycle after cadmium exposure could be suggested. Elodea canadensis does not have the anatomical feature of C4 plants, Kranz anatomy, but neither has the aquatic plant Hydrilla sp. which displays inducible C4-type photosynthesis (Reiskind et al., 1997; Rao et al., 2006). In this case, E. canadensis would have to contain enzymes for both C3 and C4 cycle enabling it to actively fix carbon dioxide under various environmental conditions.

Incorporation of bicarbonates, either by the C4 cycle or via carbonic anhydrase reaction, is known to release OH- in this species, which can be the reason for surrounding medium basification after cadmium exposure or other stress conditions e.g. DCMU or darkness (Paper I & IV). The fact that cadmium exposure resulted in increased activity of PEPC in Zea mays (Nocito et al., 2008) supports the idea of photosynthetic switching in E. canadensis. The cytosolic pH increase after cadmium exposure seems to serve as a signal for this switching by favoring the expression of PEPC (Giglioli-Guivarćh et al., 1996). Keeping in view the above mentioned facts, it is likely that E. canadensis possesses both apoplasmic and symplasmic cadmium detoxification mechanisms. Apoplasmic cadmium dextoxification depends on the special features of its cell wall. However, to clarify the mechanisms of symplasmic cadmium detoxification, further experimental work is needed.

The symplasmic cadmium detoxification may depend on calcium. Cadmium exposure 2+ induced a rise of [Ca ]cyt in E. canadensis that was cadmium concentration dependent (Paper II). The essential role of calcium in plants is well documented. It is likely that calcium is released from internal stores under cadmium exposure (Fig. 11) which ultimately helps the plant to minimize the cadmium effect on PSII (Skórzyńska-Polit and Baszyński, 2000), probably by inhibiting cadmium uptake into chloroplasts since both cadmium and calcium use the same uptake channels. Calcium is generally considered as a secondary messenger molecule in plants under various stress conditions (Kader and Lindberg, 2010). The cytosolic calcium and pH homeostasis in plant cells are closely linked (Felle, 2001). In E. canadensis, the cytosolic pH changes acts in parallel as an additional second messenger and probably contributes to the specificity of calcium signals (Paper II). The molecular mechanism of calcium-dependent heavy-metal tolerance has not yet been elucidated. A study on wheat roots, however, demonstrated the involvement of wheat root LCT1, encoding the non-specific transporter for Ca2+, Cd2+, Na+, K+ and Pb2+, in the alleviation of cadmium toxicity by calcium (Schachtman et al., 1997; Clemens et al., 1998; Antosiewicz and Hennig, 2004).

Elodea canadensis is likely to cope with cadmium exposure, apoplasmically and symplasmically, by inducing intra- and extra-cellular pH rises as shown in Fig. 11.

Figure 11. A hypothetical model summarizing the apoplasmic and cytoplasmic cadmium- detoxification mechanisms which lead to intra- and extra-cellular pH increase in E. canadensis. Cadmium exposure resulted in a significant decrease in apoplasmic pH, and cadmium binding to the cell wall was enhanced at low pH. Apoplasmic acidosis can be one of the mechanisms behind the pH rise caused by E. canadensis shoots in the surrounding medium under cadmium treatment. It might depend on the special features of its cell walls. Cadmium exposure resulted in breakage of the photosynthetic electron transport chain, inhibiting reduction of NADP+ to NADPH. Under these conditions the activity of PEPC and NADP-ME may increase. Photosynthetic isoforms of PEPC and NADP-ME may initiate C4 photosynthesis in this species after cadmium exposure. Switched photosynthesis may improve the capability of symplasmic cadmium detoxification capability of the species. Protonation of Pi, released from de-carboxylation of PEP to OAA may cause a cytosolic pH rise. Diffusion of CO2 released during de-carboxylation of malate is likely to cause cytosolic pH rise after a lag phase. Incorporation of bicarbonate, either by the C4 cycle or via carbonic anhydrase reaction, is known to release OH- in E. canadensis, which ultimately may cause the basification of the surrounding medium under cadmium treatment (or other stress conditions). Release of calcium from internal cell stores (mitochondria, ER, vacuole) under cadmium exposure may help the plant to inhibit the cadmium effect on PSII. Release of OH- may occur through some specific channel.

Root-induced basification of the growth medium by E. angustifolium after metals and arsenic exposure

The roots of E. angustifolium (common cotton grass) caused significant basification of the surrounding medium and maintained this property in the presence of different metals (copper, cadmium, lead and zinc) and a metalloid, arsenic (Paper III). The pH modulation by E. angustifolium can be a trait for its adaptation to harsh environments as E. angustifolium is commonly found in acidic wetlands. Blossfeld et al. (2010) reported rhizosphere basification by roots of alpine pennycress and ryegrass in metal contaminated soils as a permanent feature. Low concentrations of metals stimulated the root-induced basification by E. angustifolium as a maximum pH increase during the initial 24 h was observed in the presence of metals (Paper III). The release of oxalic acid, formic acid and succinic acid by E. angustifolium roots was significantly enhanced upon treatment with low levels of metals or arsenic which pointed out the significance of these acids in pH modulation after exposure to toxic elements (Paper III). Besides metals and arsenic, initial low pH can be an important factor in root-induced pH changes which ultimately increase the release of organic acids, as well as decreasing proton extrusion by plant roots. For example, heavy metal stress inhibited the activity of H+-ATPase and resulted in decreased proton efflux, causing a change in the rhizosphere pH (Astolfi et al., 2003).

At normal cytosolic pH, organic acids are present as anions in the cytosol. Therefore, the released organic acids probably take up protons when entering the rhizosphere and cause the rhizosphere pH rise after exposure to toxic elements, as observed for E. angustifolium (Paper III). The pH increase by organic acid anions, however, requires that the pH of the substrate is lower than neutral, and that the counterbalancing cation during the passage of organic acids across the cell membrane, should not be H+ (Jones and Darrah, 1994). The release of organic anions from plant cells should be balanced by the release of cations (Ryan et al., 2001). Metal toxicity to plant cells result in the loss of major nutrient cations (Kochian et al., 2005). In poplar roots, K+ ions were shown to be co-released with organic anions after metal stress (Qin et al., 2007). Metal stress may decrease the proton pumping of the H+-ATPase at the plant cell membrane. A hypothetical model describing the stimulation of organic anions from plant cell after element exposure at low pH is sketched in Fig. 12. Such mechanisms may also be working for pH increase caused by rice roots after chromium stress (Zeng et al., 2008).

Figure 12. A hypothetical model of the possible mechanisms by which free heavy metal ions cause the plant roots to release organic anions. When the substrate pH is lower than the cytoplasmic pH, the organic acids should be released as anions co-transported with essential cations (Potassium, calcium, magnesium, etc.) probably by an activation of anion channels by the metal ions. The metal stress is known to cause the loss of essential cations and inhibits the activity of H+-ATPase. These secreted organic anions may bind to protons and ultimately cause rhizosphere basification. Uptake and release of different anions and cations may occur through specific transporters.

Concluding remarks

The abilities of Elodea canadensis shoots and Eriophorum angustifolium roots to cause an increase in the pH of the surrounding medium provided the initial pH is acidic (i.e. pH 3-5), have been confirmed. The basification of the surrounding medium by these species was also observed in the presence of free toxic ions. Regarding mechanisms, the pH increase caused by E. canadensis shoots was not fully dependent on plant photosynthetic activity. This study showed that cadmium treatment caused a decrease in the apoplasmic pH of E. canadensis and resulted in cadmium binding to the cell walls. The cellular uptake of cadmium by E. canadensis follows two phases, and also contributes to the basification of the surrounding medium. Elodea canadensis initially sequesters protons and cadmium into its apoplast. The pH modulation response of E. angustifolium was prominent at low metal concentrations and partly depended on the release of oxalic acid, formic acid and succinic acid. pH-based remediation (phytofiltration and phytostabilization) may be possible by using both species, particularly when metals are present at low concentration. E.g., E. canadensis possesses cadmium detoxification mechanisms due to the presence of specific polysaccharides, presumably pectins with a low degree of esterification and glucuronic acids, in its cell walls. Eriophorum angustifolium is able to tolerate low levels of metals due to the release of organic acids.

Future prospects

The pH modulation and metal uptake response of Elodea canadensis should be investigated on multi-element polluted water to identify the relevance of this species for phytofiltration.

To further clarify the apoplasmic cadmium-detoxification, the re-organization and potential de-esterification of pectins and the proteins responsible for these processes should be the subject of future studies. Pectin methylation can be studied using specific antibodies (this work is on the way as promised by one of the collaborators). Gene expression studies for the induction of expression of genes encoding apoplasmic proteins should be performed in E. canadensis after cadmium treatment. Cadmium-induced cytosolic basification seems to be linked to symplasmic cadmium detoxification in this species, may be depending on the presence of C3 and C4 cycle enzymes. Therefore, it would be interesting to investigate the involvement of C4 cycle enzymes and their non-photosynthetic isoforms in cytosolic pH regulation and cadmium-detoxification in E. canadensis.

In Eriophorum angustifolium, the dynamics of the release of protons and essential cations at the plasmalemma after exposure to toxic ions should be studied further by using fluorescent probes.

Acknowledgements

The foremost praise goes to Almighty Allah, who enabled me to make this work done.

I am cordially grateful to Prof. Maria Greger (as my supervisor) and Prof. Birgitta Bergman (as a department head) for accepting me as a PhD student at Department of Botany, Stockholm University, Sweden. Further, I will never be able to forget the untiring efforts of my supervisor, Prof. Maria Greger, for my scientific build up and continuous support which ultimately enabled me to fulfill the tasks.

I would like to acknowledge the encouraging support of my co-supervisors Prof. Sylvia Lindberg and Prof. Birgitta Bergman. Many thanks to Prof. Sylvia Lindberg for her valuable comments on my work, as well as for being such a nice personality. Sylvia! It was always easy for me to talk with you.

I would like to extend my gratitude to all Professors and Researchers at Department of Botany (Plant Physiology Unit) for playing their important role by involving themselves in my yearly PhD evaluations. I appreciate the efforts of Prof. Katharina Pawlowski and Prof. Lisbeth Jonsson to improve this thesis with constructive comments. It will be pity if I do not compliment the impressive way of Prof. Peter Hambäck of handling peoples and their issues.

I would like to acknowledge Dr. Britta Eklund, Dr. Anna Ohlsson and Dr. Mats Björk for evaluating my licentiate thesis. The critical analysis of my licentiate thesis by Prof. Mats Björk was constructive and opened new ways of thinking for me.

All the members of “Plant Metal Group” Claes (being an indispensable roommate), Clara, Johanna, Tommy, Sandhi, Prem, Kathrin, Agneta, Beata, Lillivor, Lukas, Daniel are acknowledged for all interesting fikas, summer lunches and fun discussions. Your presence makes my stay here very comfortable. My special thanks to Tommy Landberg, for assisting me in all the laboratory matters, and sparing time to answer all my questions. Thanks to Peter and Ingela for being always supportive in the green house.

My studies at Stockholm University were the part of a PhD studentship from Higher Education (HEC) Pakistan administrated by the Swedish Institute (SI). Research funding was provided by the Carl Tryggers foundation and the Knut and Alice Wallenberg foundation, Sweden. Financial means for international courses and conferences were provided by the Department of Botany, Stockholm University, the Royal Swedish Academy of Sciences (KVA), and the Royal Swedish Academy of Agriculture and Forestry (KSLA).

Dr. Basra and Dr. Afzal (Department of Crop Physiology, University of Agriculture, Faisalabad, Pakistan) deserve immense thanks at this milestone of my academic carrier as they introduced me to plant stress physiology.

Outside academia, my thanks go to my family, friends and relatives-thank you all to keep me up.

References

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