Evaluation of Australian Native Wetland Plants for Phytoremediation of Saline Mine–Leachate

Evaluation of Australian native wetland plants for phytoremediation of saline mine–leachate

Allan Andrew Adams B. Ecol. Agric. (Hons) (Univ. Sydney)

A thesis submitted for the fulfilment of the requirements for the degree of

Doctor of Philosophy

Faculty of Science

School of Agricultural and Wine Sciences

Table of Contents

List of Figures ...... 5 List of Tables ...... 7 Certificate of Authorship ...... Error! Bookmark not defined. Preface ...... 12 Acknowledgements ...... 13 Publications from Thesis ...... 14 Abstract ...... 15 Chapter 1: Introduction ...... 17 1.0 Introduction ...... 17

1.1 Generation of heavy metal leachate due to mining activities ...... 20

1.2 Environmental consequences of mine leachate ...... 22

1.3 Phytoremediation ...... 24

1.4 Phytoremediation of heavy metals in mine wastes ...... 27

1.5 Heavy metal accumulation and levels of tolerance by plants ...... 34

1.6 Physiological mechanisms in phytoremediation ...... 36

1.7 Genetic engineering for enhanced heavy-metal storage in accumulating and tolerating plants ...... 39

1.8. Efficacy of wetland plants in remediating heavy-metal-containing wastewater ...... 41

1.8.1 Wetland species selection for phytoremediation ...... 48

1.8.2 Effects of heavy metals on plants ...... 51

1.9 Metal removal processes in wetlands ...... 52

1.9.1 Wetland plant removal of metals ...... 52

1.9.2 Physical removal of metals ...... 53

1.9.3 Chemical removal of metals ...... 53

2.0 Cadia Valley Operations ...... 54

2.1 Waste rock dumps at Cadia Valley Operations ...... 56

2.2 Aims and objectives ...... 58

Chapter 2: Accumulation of heavy metals by naturally colonising Typha domingensis (Poales: Typhaceae) in waste-rock dump leachate storage ponds in a gold–copper mine in the central tablelands of New South Wales, Australia ...... 63 2.1 Introduction ...... 63

2.2 Materials and methods ...... 66

2.2.1 Study sites ...... 66

2.2.2 Sample preparation and analysis ...... 69

2.3.3 Data analysis ...... 71

2.4. Results ...... 72

2.4.1 Leachate and creek water analysis ...... 72

2.5 Sediment analysis (DTPA extraction, 0–5 cm and 5–15 cm depths) ...... 73

2.6 Sediment analysis (Total acid extraction, 0–5 cm and 5–15 cm depths) ...... 74

2.7 Plant analysis (Roots and shoots) ...... 74

2.8 Translocation factor ...... 77

2.9. Discussion ...... 77

Chapter 3: Hydroponic assessment of four emergent wetland species for phytoremediation of wetlands impacted by saline-mine leachate ...... 83 3.1 Introduction ...... 83

3.2 Materials and methods ...... 85

3.2.1 Plant material and treatments ...... 85

3.2.2 Determination of porewater reference values for hydroponic solution ...... 86

3.2.3 Growth measurements ...... 88

3.2.4 Gas-exchange measurements ...... 88

3.2.5 Chlorophyll-fluorescence measurements ...... 89

3.2.6 Element accumulation ...... 89

3.2.7 Statistical analysis ...... 90 3

3.3 Results ...... 90

3.3.1 Growth measurements ...... 90

3.3.2 Gas-exchange measurements ...... 94

3.3.3 Chlorophyll fluorescence: Fv/Fm, NPQ, and PhiPS2 ...... 95

3.3.4 Metal accumulation ...... 96

3.4 Discussion ...... 101

Chapter 4: Carbon assimilation and porewater interactions of macrophytes for phytoremediation of saline mine leachate ...... 106 4.1 Introduction ...... 106

4.2 Statistical analysis ...... 109

4.3 Material and methods ...... 110

4.3.1 Plant materials and treatment ...... 111

4.3.2 Particle size assessment ...... 112

4.3.3 Mine waste-rock dump leachate ...... 113

4.4.1 Carbon assimilation ...... 113

4.4.1.1 Gas exchange measurements ...... 113

4.4.1.2 Chlorophyll fluorescence measurements ...... 114

4.3.1.3 Leaf tissue preparations for carbon and nitrogen isotope ratio measurements ...... 114

4.4.1.4 Carbon and nitrogen concentration determinations and isotope ratio measurements ...... 115

4.4.1.5 Measurement of Rubisco Levels ...... 115

4.4.1.6 Preparation of leaf-tissue extracts ...... 116

4.4.1.7 Determination of Total Soluble Protein: ...... 116

4.4.1.8 Rubisco quantitation ...... 116

4.4.1.9 Rubisco Standard Curve: ...... 117

4.5 Mineral accumulation analysis ...... 117

4.6 Porewater collection ...... 118 4

4.7 Growth measurements ...... 119

4.8 Results ...... 119

4.8.1 Carbon assimilation ...... 119

4.8.1.1 Gas exchange measurements ...... 119

4.8.1.2 Carbon and nitrogen leaf percentages ...... 123

4.8.1.3 Carbon and Nitrogen isotope ratios ...... 123

4.8.1.4 Rubisco quantitation ...... 124

4.8.1.5 Gas exchange relationships ...... 125

4.9 Mineral accumulation analysis ...... 127

4.9.1 Plant growth measurements...... 130

4.9.2 Bioconcentration and translocation ...... 131

4.9.3 Elements in porewater ...... 135

4.8 Discussion ...... 142

Chapter 5: General conclusion ...... 152 5.1 Summary of experimental aims ...... 152

5.2 Plant evaluation ...... 152

5.2 Concentration of elements in substrates and root-zone interactions ...... 154

5.3 Photosynthetic performance ...... 154

5.4 Accumulation and translocation of elements ...... 155

5.5 Conclusions ...... 156

List of Figures Figure 1. Ficklin diagram showing acid-rock drainage, neutral mine drainage, saline drainage (Plumlee et al. 1999)...... 46

Figure 2. Aerial image of Cadia Valley Operations, Orange, NSW (Scale: 1 cm = 0.5 km)...... 55

Figure 3. Arial image of sampling points (dots) within T. domingensis stand in NLP (northern leachate pond) at the foot of the SWRD (outlined in black) (Scale: 1 cm = 23 m)...... 69

Figure 4. Aerial image of mine site showing the open pit, south waste-rock dump, and tailings dam. CAC - Cadiangullong Ck, NLP - northern leachate pond, SLP - southern leachate pond, SWRD - south waste rock dump (Scale: 1 cm = 0.5km)...... 70

Figure 5. a-c: Curvilinear regression of Cu (a), Mn (b), and Zn (c) concentrations (mg/L) from leachate and creek-water sampling points (1-5) on the 50 m transect at 10 m intervals between each sampling point in early autumn (13-16 March 2011) at northern leachate pond (NLP) n=6...... 73

Figure 6. A-B. Images of T. domingensis roots collected from (A) CAC (Cadiangullong Creek) showing Fe-root plaque (Bar = 1.2cm) and (B) NLP (northern leachate pond) with absence of Fe-root plaque (bar = 1 cm)...... 75

Figure 7. The effect of treatments (Conc. 1–4) against control (0) in the hydroponic medium on root length (a), dry mass (b) of B. articulata, C. appressa, P. australis, and S. validus. Bar indicates l.s.d. at 5% level, n=6...... 91

Figure 8. The effect of treatment (Conc. 1–4) against control (0) in the simulated leachate solutions on photosynthetic rate (A), stomatal conductance (B), internal CO2 concentration (C), and transpiration rate (D). Bar indicates l.s.d. at 5 % level, n=6...... 94

Figure 9. The effect of treatment in the hydroponic medium on Fv/Fm (A), PhiPS2 (B), and Non-photo chemical quenching (NPQ) (C). Bar indicates l.s.d. at 5 % level, n=6...... 96

Figure 10. Photosynthetic rate (a), stomatal conductance (b), internal CO2 concentration (c), and transpiration rate (d) of B. articulata (B. a), C. appressa (C. a), P. australis (P. a), and S. validus (S. v) grown in wetland sediment impacted by saline mine leachate. Mean data presented with standard deviation, columns with different letters (superscript) indicate significant difference between control (white columns) and treatment (black columns) at l.s.d at 5 % level, n=8...... 121

Figure 11. Figure 2. Maximum quantum efficiency of PSII – Fv/Fm (a), PhiPS2 (b), and Non-photo chemical quenching (NPQ) (c) of B. articulata (B .a), C. appressa (C. a), P. australis (P. a), and S. validus (S. v) grown in wetland sediment impacted by saline mine leachate. Mean data presented with standard deviation, columns with different letters (superscript) indicate significant difference between control (white column) treatment (black column) at l.s.d at 5 % level, n=8...... 122

Figure 12. Immunoblot showing the different amounts of Rubisco (large subunit) in the different tissue extracts of B. articulata, C. appressa, P. australis, and S. validus when identical amounts of soluble protein were loaded on the gel; C = control and T = treatment for each species (a), immunoblot showing Rubisco used for constructing the standard curve (b)...... 125 6

Figure 13. Impact of Mg, Mn, and S in porewater on photosynthetic performance (a,b,c), Mg and Na in porewater on photo chemical quenching (NPQ) (d,e), and Cu, Mn, and Zn in porewater on freshweight (milligrams), (f,g,h) on the selected species...... 127

Figure 14. Image showing pots with Rhizon porewater samplers inserted into the side of each pot...... 130

Figure 15. Fresh (a) and dry (b) mass of B. articulata (B.a), C. appressa (C.a), P. australis (P.a), and S. validus (S.v). Mean data presented with standard deviation, columns with different letters indicate significant difference between control (white columns) and treatment (black columns) at l.s.d at 5 % level, n=6...... 131

Figure 16. Mean concentrations (mg/L) of porewater extracted from 4 L plastic pots containing control and treatment wetland sediment. Figure 15. a–(Ca), b–(Cu), c–(Mg), d– (Mn), e–(Na), f–(S), g–(Zn), h (mineral N), i (ORP), j (pH), h (Temp). B.a C = B. articulata (Control), B.a T = B. articulata (Treatment), C.a C = C. appressa (Control), C.a T = C. appressa (Treatment), P.a C = P. australis (Control), P.a T = P. australis (treatment), S.v C = S. validus (Control), S.v T = S. validus (treatment). S1, S2, and S3 indicate the three sampling periods over the 8 month experimental period. Columns with different letters (superscript) indicate a significant difference, significance value was set at (P<0.05). Standard deviation indicated by rods within columns. Log to base 10 values of ORP, Mineral N, and Mn less than 0.1 will be negative, graphs shown on log scale, n=8...... 141 List of Tables Table 1. Metal concentrations ppm in igneous and sedimentary rocks (Cannon et al. 1978)...... 22

Table 2. Comparison of conventional and phytoremediation methods for treatment of heavy metals ...... 31

Table 3. Accumulation of heavy metals by wetland plants in field, glasshouse and hydroponic trials...... 32

Table 4. Wetland treatment of mine drainage from CMA mine in Tucush Valley waste dump (Estimated parameter concentration reduction - based on average loading rates) (Strachotta et al. 2009)...... 47

Table 5. Metal-removal rates (mg m-2 day-1) from pilot-scale surface-flow wetlands at Tara Mines, Navan, Ireland calculated from differences between water entering and water exiting the treatment system as a function of the average flow rate (1.5 L min-1), n=5. Equivalent percentage removals are also included, - implies net export of metal (O’Sullivan et al. 2004)...... 47

Table 6. Mean values of Cu, Mn, and Zn (mg/kg) in sediment from DTPA-extractable and total (acid digested) at 0-5 cm and 5-15 cm depths and root and shoots of T. domingensis at

northern leachate pond (NLP), southern leachate pond (SLP), and Cadiangullong Ck (CAC) during early autumn (13-16 March, 2011) and winter (11-14 July, 2010) sampling, n=6. .... 76

Table 7. Translocation factors in T. domingensis at Cadiangullong Creek (CAC), northern leachate pond (NLP), and southern leachate pond (SLP)...... 77

Table 8. Porewater concentrations (mg/L) of Ca, Cu, Mg, Mn, S, and Zn from leachate storage pond receiving saline-mine leachate from a gold and copper mine in Central- tablelands, NSW, Australia (Adams et al. 2013). Mean data and SD values presented, n=6.87

Table 9. Concentrations of elements dissolved in hydroponic solution...... 87

Table 10. Tolerance index % of B. articulata, C. appressa, P. australis, and S. validus. Mean data presented and standard deviation, numbers with different letters indicate significant difference within treatment at l.s.d at 5 % level...... 91

Table 11. Means and standard deviations for bioconcentration factor (BCF) (Table 12 A) and translocation factor (TF) (Table 12 B) of B. articulata, C. appressa, P. australis, and S. validus.” The same letter of alphabet, superscripted, within columns means the compared data do not differ significantly (p<0.05). Use of ‘n/a’ against data pertaining to P. australis was done because of insufficient leaf material to carry out analysis in triplicate...... 92

Table 12. Means and standard deviations for concentrations (mg/kg) of Ca, Cu, Mg, Mn, S, and Zn in shoots and roots of B. articulata, C. appressa, P. australis, and S. validus”. The same letter of the alphabet, superscripted, within columns means the compared data do not differ significantly (p<0.05) between treatments and species, no letters indicate no significant difference, n=6...... 98

Table 13. Concentrations of elements in acid- mine drainage and saline mine-drainage (Plumlee et al. 1999)...... 108

Table 14. Chemical composition (mg/L) of mine waste-rock dump leachate and control pond water collected during study...... 112

Table 15. Particle size analysis of sediment collected from control and treatment pond. ... 113

Table 16. Carbon and nitrogen percentages in leaf material of B. articulata, C. appressa, P. australis, and S. validus grown in wetland sediment impacted by saline leachate. Mean data presented with standard deviation, different letters indicate significant diff different letters indicate significant different letters indicate significant difference between control and treatment at l.s.d at 5 % level, n=5...... 123

Table 17. Signatures of carbon and nitrogen isotopes in leaf material of B. articulata, C. appressa, P. australis, and S. validus grown in wetland sediment impacted by saline leachate. Mean data presented with standard deviation, different letters indicate significant difference between control and treatment at l.s.d at 5 % level, n=5...... 124

Table 18. Rubisco levels (nmol) per mg of total soluble leaf protein in B. articulata, C. appressa, P. australis, and S. validus...... 125

Table 19. Mean concentration and standard deviation of mineral accumulation in roots and shoots of B. articulata, C. appressa, P australis, and S. validus grown in wetland sediment impacted by saline mine leachate. Mean data presented with standard, different letters (superscript) indicate significant difference between control and treatment at l.s.d at 5 % level, n=6...... 129

Table 20. Mean bioconcentration factor (a), and translocation factor (b) of B. articulata, C. appressa, P. australis, and S. validus and standard deviation. Values with different letters (superscript) indicate significant difference between treatment and control at l.s.d at 5 % level, n=6...... 133

List of Acronyms

AMD Acid mine drainage ANOVA Analysis of variance ARD Acid rock drainage ATP Adenosine 5’ triphosphate BCF Bioconcentration factor CAC Cadiangullong Creek CaMV Cauliflower mosaic virus CVO Cadia valley operations CWs Constructed Wetlands DM Dry mass DTPA Diethylene triamine pentaacetic acid DMSe Dimethyl selenide EDTA Ethylene diamine tetra acetic acid GRS Green remediation strategy GS Glutathione synthetase LCF Leaf chamber fluorescence MerB Organomercurial lyase MT Metallothionein NAF Non acid forming NMD Neutral mine drainage NLP Northern leachate pond NPQ Non photo chemical quenching PAF Potentially acid forming PVP Poly vinyl pyrrolidine PsMTA Pisum sativum mitochondrial ROL Radial oxygen loss ROS Reactive oxygen species SeCys Selenoaminoacids selenocysteine SeMet Seleno methionine SLP Southern leachate pond SMT Selenocysteine methyltransferase SWRD South waste rock dump TEA Triethanolamine TF Translocation factor TI Tolerance index TNT Tri nitro toluene TSF Tailings storage facilities USEPA United States Environment Protection Agency UV Ultra Violet

Certificate of Authorship

I hereby declare that this submission is my own work and to the best of my knowledge and belief, understand that it contains material previously published or written by another person, no material which to a substantial extent has been accepted for the award of any other degree or diploma at Charles Sturt University or any other educational Institution, except where due acknowledgement is made in the thesis [or dissertation, as appropriate]. Any contribution made to the research by colleagues with whom I have worked at Charles Sturt University or elsewhere during my candidature is fully acknowledged,

1 agree that this thesis be accessible for the purpose of study and research in accordance with normal conditions established by the Executive Director, Library Services, Charles Sturt University or nominee, for the care, loan and reproduction of thesis, subject to confidentiality provisions as approved by the University,

Name Allan Adams

Date 14/11/2014

Preface

The primary aim of this thesis was to determine if several selected wetland plants were capable of tolerating the conditions of a pond/wetland system receiving leachate from a mine waste-rock dump at a Au and Cu mine - Cadia Valley

Operations (CVO). The present study focuses on the selection and evaluation of several Australian native wetland plant species for phytoremediation of saline mine- leachate. The study was conducted within the mining lease of (CVO), owned and operated by Newcrest Mining Pty Ltd and within glasshouse facilities at Charles

Sturt University, Orange, NSW.

At the commencement of the project several key objectives were determined, (1) assess the performance of Typha domingensis, this species had naturally colonised many of the leachate and water storage ponds at CVO, and visually appeared to be tolerant of the conditions, (2) classify the type of leachate generated within the waste-rock dumps and the excess concentrations of elements within the pond/wetland systems, (3) select several Australian native wetland plant species suitable for trials within the wetland conditions receiving the waste-rock dump leachate. The study was conducted over 3.5 years and included experiments in laboratory, glasshouse and field conditions.

Acknowledgements

I thank my supervisors Anantanarayanan Raman and Dennis Hodgkins for their continued support throughout my post-graduate study.

I also thank my partner Jos and Ryan for allowing me to work the many nights and weekends to complete my studies.

My project was funded by the Australian Post Graduate Award, Rural Management Fund, and Cadia Valley Operations; I greatly appreciate these funding sources.

I would also like to acknowledge the wealth of knowledge and assistance from David Wade (EAL – CSU). I also thank Madhavan Soundararajan (University of Nebraska- Lincoln) for enabling me to travel to his laboratory and for his assistance and advice with experimental work. I also thank Lesley Batty (University of Birmingham) for the endless responses to my many emails.

In addition, I would like to thank Karen Gogala, Jeff West and all the staff at Charles Sturt University that have assisted me over my years of study.

I am also greatly appreciative of Helen Nicol for her advice with matters of Statistics, and the environment team at Cadia Valley Operations, in particular, John Ford and Jeff Burton.

This thesis is a zenith of inspiration and search for answers that was fostered in me by my parents; this thesis is dedicated to my mum and dad.

Publications from Thesis

Adams AA, Raman A, Hodgkins DS, Nicol N. (2013) Accumulation of heavy metals by naturally colonising Typha domingensis (Poales: Typhaceae) in waste-rock dump leachate storage ponds in a gold and copper mine in the central tablelands of New South Wales.

International Journal of Mining, Reclamation, and Environment 27: 294–307.

Adams A, Raman A, Hodgkins D. (2013) How do the plants used in constructed wetlands, a sustainable remediation strategy, perform in heavy-metal-contaminated mine sites. Water and Environment Journal 27: 373–386.

Abstract

The aim of this thesis was to assess the suitability of the Australian native wetland plants Typha domingensis, Baumea articulata, Carex appressa, Phragmites australis, and Schoenoplectus validus for use in wetland phytoremediation of saline mine-leachate. Experimental work was conducted in glasshouse and laboratory conditions, at an active Au and Cu mine in the Central Tablelands of NSW,

Australia. The evaluation of each species consisted of three main themes, (1) concentration of elements in wetland substrates and root-zone interactions, (2) photosynthetic performance including chlorophyll fluorescence and determine of

Rubisco concentrations, and (3) accumulation and translocation of Ca, Cu, Mg, Mn,

Na, S, and Zn.

Naturally colonised population of Typha domingensis in artificial wetlands were found to assist in removal of Cu, Mn, and Zn from saline mine-leachate exiting waste-rock dumps at the mine site. Hydroponic trials were used to screen the suitability of B. articulata, C. appressa, P. australis, and S. validus by determining their element accumulation and photosynthetic performance; the findings from this experiment suggested their suitability for further experimental trials. Pot experiments containing wetland sediment and soil were used to determine carbon assimilation performance; a significant difference in non-photochemical quenching (NPQ) of control and treatment populations of B. articulata occurred, greater amounts of

Rubisco were found in treatment populations of B. articulata, C. appressa, P. australis, and S. validus. Pot trials also revealed fresh mass of B. articulata and C. 15

appressa were significantly reduced and dry mass of C. appressa only were significantly reduced. Carex appressa accumulated the greatest amount of Ca, Cu,

Mn, Na, S and Zn, and also translocated the greatest amounts of ca, Cu, Mg, Na, and

S. Assessment of porewater with Rhizon samplers revealed significant increases of

Cu, Mg, and S in treatment porewater, and concentrations of Mn were significantly decreased in the presence of P. australis in treatment conditions only.

Chapter 1: Introduction

1.0 Introduction

Mining today has grown substantively because of the rising value of metals, such as copper (Cu), gold (Au), nickel (Ni), silver (Ag), lead (Pb) and zinc (Zn). Au production in Australia alone has risen from 17 035 t in 1980 to 244 472 t in 2007

(Mudd 2009). Although this rise brings economic advantages, the environmental downside is that during metal-ore extraction and purification, large volumes of heavy-metal waste generation occurs, which may ultimately be discharged into the environment via surface drainage and groundwater recharge. Between 1950 and

2000, 22 000 t of cadmium (Cd), 939 000 t of Cu, 783 000 t of Pb and 1 350 000 t of

Zn have been discharged into aquatic environments in different parts of the world

(Singh et al. 2003). Globally, because of direct mining operations, 37 000 km2

(Dudka & Adriano 1997) and a surrounding 240 000 km2 have been affected by heavy-metal discharge from mine waste rock dumps and metal-including drainage

(the leachate) (Furrer et al. 2002). High levels of Cd, Cu, Cu, Pb and iron (Fe), and their salts occurring in the leachate can be toxic to organisms (Guilizzoni 1991).

Being principal-soil contaminants, the heavy metals are the most difficult and expensive to remove (Ensley 2000). This is of concern because of the high levels of toxicity that can eventuate even at low concentrations of metals and their tendency to bioaccumulate. For instance, metal accumulation in aquatic sediments, because of mining in Northern Idaho, induced inflammation in and necrosis of tissue in the edible bull trout (Salvelinus confluentus) (Kiser et al. 2010). Moreover, long-term 17

exposure to heavy metals [e.g., Fe, Zn, Cu and manganese (Mn)] and metalloids

[e.g., arsenic (As)] in humans are known to induce neurotoxicity resulting in neurobehavioral disorders (Newschaffer et al. 2007; Wright & Baccarelli 2007).

Heavy metals can be removed from wastewater using physical and chemical methods: oxidation, reduction, precipitation, membrane filtration, ion exchange, electrochemical operation, biological treatment and adsorption (Corami et al. 2007).

These methods can induce adverse effects on the biota, structure and fertility of the soil (Mulligan et al. 2001). The techniques used in soil remediation can be utilised in treating sediments in wetlands in mine sites. Further to physical and chemical methods, soil remediation efforts employ civil engineering (e.g., capping, excavation), thermal (e.g., electrical resistance heating, steam injection and heating) and biological (e.g., bioremediation, phytoremediation, chemical-assisted phytoremediation) techniques. Biological remediation, a less expensive and a more sustainable option, involves the use of plants to either remove or sequester hazardous substances (Gräfe & Naidu 2008). Chemical-assisted bioaugmentation is also used currently employing synthetic chelating agents, such as Ethylenediaminetetraacetic acid (EDTA), in combination with plants (Lombi et al. 2001). Between the late

1980s and early 1990s, bioremediation to treat hydrocarbon wastes using chemicals in conjunction with the application of bacteria (e.g., Bacillus subtilis, Pseudomonas aeruginosa) for groundwater and surface-water remediation was popular. Several in situ bioremediation strategies have been tested, of which at least 10% have been found to degrade halogenated aliphatic compounds released into aquatic environments (Major & Cox 1992).

Excavation of contaminated sediments and storing them at nominated sites are customarily done in Australia. However, aiming at measures of sustainability aligned with natural management, the current federal- and state government regulations urge using sophisticated on-site management tactics rather than using traditional practice of civil work-based technology. Civil work-based technology bears several environmental disadvantages: exhaustion of natural resources and energy, generation of large volumes of waste, and emission of CO2 and diesel fumes. Use of large quantities of chelators [e.g., EDTA, diethylene triamine pentaacetic acid (DTPA)] alters the structure and damages soil, its texture and fertility, and can lead to contamination caused by mobile heavy-metal ions leaching into groundwater. The leaching mobile heavy-metal ions are nonbiodegradable (Wu et al. 2010). The

Superfund Green Remediation Strategy (GRS), launched by United States

Environmental Protection Agency (USEPA) (2009), recommends reducing the exhaustion of natural resources, energy, greenhouse gas emissions during remediation of hazardous waste dump sites. Optimisation of energy-intensive systems results in substantial reductions in natural resource and energy consumption.

Remediation technologies used in treating organic and inorganic contaminants at mine sites such as pumping and treating, thermal desorption, multiphase extraction, air sparging, and soil-vapour extraction simulate plants used in decontamination activities. Thermal treatments release metals in flue gases, while incineration and soil-vapour extraction lead to vapour emission (Castelo-Grande & Barbosa 2003).

Plant based remediation efforts, in contrast, are a low-energy demanding alternative because plants sequester heavy metals and inorganic complexes through

accumulation and stabilisation. While comparing the potential environmental footprint of phytoremediation with the alternative treatment technologies, such as chemical treatments, phytoremediation impresses as the sustainable alternative

(Glass 2000).

Effective and sustainable management of leachate is of high priority in environmental management at mine sites. One such option for heavy-metal removal from leachate is the use of constructed wetlands (CWs) (Ernst 2005). CWs at mine sites are suitable for remediation of a variety of water quality issues (Vymazal &

Kröpfelová 2008). CWs are designed structures that mimic natural wetlands in terms of their sedimentation and biological (microbes and plants) functions (Rai 2008).

Wetlands increase the retention time of contaminants by reducing water flow, therefore minimising the diffusion of pollutants into the surrounding environment

(An et al. 2011). Heavy metals are accumulated in surface waters, groundwater, substrates and plants (Aksoy et al. 2005). Use of CWs at mine sites is a low-energy alternative for removing metals from the leachate and retaining the metals in sediments instable forms.

1.1 Generation of heavy metal leachate due to mining activities

Leachate is generated within waste-rock dumps at mine sites by oxidation of sulfide minerals, usually pyrites (FeS2). Pyrite oxidation involves chemical, biological, and electrochemical reactions and varies with environmental conditions. The rate of

oxidation is determined by pH, PO2, specific surface structure of FeS2, presence and/or absence of bacteria (e.g., Ferrobacillus ferrooxidans) and clay minerals, and hydrological factors (Evangelou and Zhang 1995). Oxidation of FeS2 can produce acid-rock drainage (ARD) with a pH <6, neutral-mine drainage (NMD), and saline drainage pH >6 (Plumlee at al. 1999). The liquid contains dissolved levels of any constituents that are solubilised, and depends on the chemical properties of the waste material and water chemistry (Halford 1999). The composition and concentration of heavy metals depend on the geological-parent material (Table 1) and environmental conditions that activate the weathering process (Nagajyoti et al. 2010).

Pyrite oxidation process:

2+ 2– + FeS2 + 7/202 + H2O Fe + 2SO 4 + 2H (1)

2+ + 3+ Fe + 1/4O2 + H Fe + 1/2H2O (2)

3+ + Fe + 3H2O Fe (OH)3 (s) + 3H (3)

FeS2 + 7Fe2 (SO4)3 +8H2O 15FeSO4 + 8H2SO4 (4)

2+ Reaction 1 shows the reaction process of O2 with FeS2 to produce Fe , which is

3+ oxidised to Fe by O2 (2). Reaction 3 shows formation of Fe(OH)3 precipitate

(Singer and Stumm 1970). In the presence of Thiobacillus ferrooxidans and in pH below 7, pyrite oxidation occurs (2, 4) (Evangelou and Zhang 1995). Oxidation of pyrite and subsequent production of leachate not only contaminates natural waters with sulphate, iron, and acidity, but also solubilises heavy-metals (e.g., Cu, Zn).

Table 1. Metal concentrations ppm in igneous and sedimentary rocks (Cannon et al. 1978).

Basaltic igneous Granite igneous Shales and Clays Black shales Sand stone As 0.2–10 0.2–13.8 – – 0.6–9.7 Cd 0.006–0.6 0.003–0.18 0.0–11 <0.3–8.4 – Cr 40–600 2–90 30–590 26–1,00 – Co 24–90 1–15 5–25 7–100 – Cu 30–160 4–30 18–120 20–200 – Pb 2–18 6–30 16–50 7–15000 <1–31 Mo 0.9–7 1–6 – 1–300 – Ni 45–410 2–20 20–250 10–500 – Zn 48–240 5–140 18–180 34–1,500 2–41

Copper smelting during Roman and Medieval times was a turning point in the metal

‘industry’ in causing widespread pollution (Hong et al. 1996). Heavy metals in soils and waterways have increased considerably since the Industrial Revolution, reaching toxic levels in industrial and mining regions throughout the world (Rybicka 1996).

The annual worldwide release of heavy metals reached 22, 000 t for Cd, 939, 000 t for Cu, 783, 000 t for Pb, and 1, 350, 000 t for Zn (Singh et al. 2003). Heavy metals contaminate soil extensively and are the most difficult and expensive to remove. In the United States alone more than 50, 000 metal contaminated sites await removal

(Ensley 2000). Estimated costing to clean up one acre of heavy-metal contaminated soil using plants is US $ 60, 000–100, 000, whereas alternative remediation using physical-soil excavation is at least US $ 4,000,000 (Salt et al. 1995).

1.2 Environmental consequences of mine leachate

Controlled and uncontrolled disposal of mining and smelting metalliferous ores results in contaminants entering non-contaminated sites as either dust or leachate and 22

contaminate landscapes (Ghosh and Singh 2005). The negative effects of heavy metals include reduced agricultural productivity, contaminated food sources, and cancer, respiratory diseases in humans. Heavy metals are also transferred to mammalian recipients through the food chain (Brams et al. 1989; Davydova 1999;

Hall 2002). Given the potential adverse impacts on environment and human health, a global effort has focused on the remediation of contaminated sites, with human communities demanding cleanup of such sites (Gräfe and Naidu 2008).

Stripping mountain tops to expose coal seams is common practice in Kentucky, West

Virginia and south-western Virginia, USA. Forests are stripped and rock is broken up with explosives and the rubble is dumped into valleys often filling streams

(Gilbert 2010). The water from the debris includes heavy metals and sulphates, with an electrical conductivity at 1,100 µS cm 1, whereas the maximum allowable concentration recommend by the US EPA is 500 µS cm.1; decreases in stream invertebrate biodiversity have been correlated with increased sulphate and metal concentrations (Gilbert 2010). Metal contamination due to mining activities in waterways is covered further in chapter 2.

Mining is a major contributor to Australian economy. Australia includes numerous mine sites, which have left several critical environmental legacies impacting on surrounding and downstream ecosystems due to heavy metal discharge and accumulation. The Ag, Au, Cu mine at Mt Lyell, Au, Cu at Mt Morgan and the U,

Cu project at Rum Jungle have caused significant environmental problems in the

surrounding water ways (Mudd 2009). Rum Jungle site was infamous for its pollution; approximately 1 million L/day of liquid effluent was discharged into the

Finniss River, containing acids (pH 1.5), metals and radionuclides, resulting in the disappearance of biota downstream and contamination of the floodplains – 100 km2

(Davy 1975).

1.3 Phytoremediation

The extensive research outcomes have led many to refer to phytoremediation as a

‘technology’ (phytotechnology) (Lombi et al. 2001; Arthur et al. 2005; Ernst 2005;

Pilon-Smits 2005; Dickinson et al. 2009). This technology for environmental restoration has increasingly split into various types, the main division being between phytoextraction and phytostabilisation (Peer at al. 2005).

Phytoextraction utilises plant species known to accumulate metals and produce high biomass, metals are transported and concentrated into harvestable roots and above- ground shoots (Lasat 2002). Those plant species known to accumulate large quantities of heavy metals are the hyperaccumulators (Chaney et al. 1997). Plant species capable of accumulating metals in the shoots at levels > 1.0% for Zn and Mn,

0.1% for Al, As, Se, Ni, Co, Cr, Cu, and Pb, and 0.01% for Cd are hyperaccumulators (Baker and Brooks 1989). Hyperaccumulation capacity is known from approximately 400 species of flowering plants, which take up, transport and sequester metals, achieving tissue concentrations that are otherwise toxic to most

organisms (Baker at al. 2000). The immediate practical application of these species is to either cleanse contaminated soil (Cunningham and Ow 1996), or use such plants to extract commercially valuable metals from low–grade ores for phytomining

(Brooks et al. 1998). Also known as phytoaccumulation, phytoextraction removes contamination from soil without destroying the soil structure and fertility (Ghosh and

Singh 2005). This technique is best suited to remediation of areas where pollutants occur at relatively low concentrations (Rulkens et al. 1998). The two prominent approaches of phytoextraction are (a) chelate assisted extraction or induced extraction, artificial chelates are added to increase the mobility and uptake of metal contaminant, and (b) continuous extraction, although dependant on the ability of the plant to remediate, only plant growth repetitions are controlled (Salt et al. 1995).

Phytostabilisation is the reduction of the flow of one or more contaminants into the environment. Plants capable of tolerating high metal concentrations provide a stable soil cover and they also restrict contaminants moving into groundwater or air

(Tordoff et al. 2000). Phytostabilisation is usually used when phytoextraction is neither possible nor desirable and aims to reduce the flow of contaminants in the environment. Certain plants immobilise contaminants in soil through absorption and accumulation by roots, adsorption onto roots, and/or precipitation in the root zone along with physical stabilisation of soils (Padmavathiamma and Li 2007). Vegetation capable of tolerating high-metal concentrations provide a soil cover in conditions unsuitable for non-metal tolerant species, which reduces the mobility of contaminants and prevents migration to either groundwater or air (Tordoff et al.

2000). The use of lime, phosphate and compost amendments to assist phytostabilisation is an effective way to reduce the bioavailability of metals in soil and uptake by plants (Adriano et al. 2004; Ruttens et al. 2006). The major disadvantage with phytostabilisation is that the contaminant remains in the medium

(Ghosh and Singh 2005).

Each process involves various mechanisms for the decontamination of pollutants from the environment; all such processes either decontaminate the soil or stabilise the pollutant within it (Singh et al. 2003); this includes phytoextraction, phytostabilisation, phytovolatilisation, phytofiltration, phytomining.

 Phytoextraction — accumulation of metals by plant species known to produce

high-biomass, metals are transported and concentrated into harvestable roots and

above-ground shoots (Lasat 2002).

 Phytostabilisation — reduction of the flow of one or more contaminants into the

environment. Plants capable of tolerating high metal concentrations provide a

stable soil cover and they also restrict contaminants moving into groundwater or

air (Tordoff et al. 2000).

 Phytovolatilisation — transmission of contaminants through the stomata into the

atmosphere as a gas (Pilon-Smits 2005), i.e., inorganic Se can be volatilised by

assimilation into the organic selenoaminoacids selenocysteine (SeCys) and

selenomethionine (SeMet).

 Phytofiltration — use of plants as filters in wetlands (Horne 2000).

 Phytomining — use of plants in extracting commercially valuable metals from

low–grade ores (Brooks et al. 1998).

1.4 Phytoremediation of heavy metals in mine wastes

Uptake and accumulation of heavy metals directly from water bodies and assimilation by plants are the greatest benefit of phytoremediation (Rai 2009).

Phytoremediation is a low-cost, sustainable and green remediation strategy for cleansing heavy metals in contaminated landscapes (Arthur et al. 2005).

Phytoremediation offers a turnkey mechanism for sustainability during mining and at mine closure. As a solar driven technique, building on the innate capabilities of plants in remediating soil and water environments in the least disturbing manner, it is a clean production technique (Dickinson et al. 2009). It is sustainable because it meets the economic, sociological and environmental needs of not only the present human generation but also those of a foreseeable number of future generations.

Nevertheless, critical challenges do exist: the efficacy of phytoremediation is debatable when contaminants occur at depths inaccessible to plant roots (Keller et al.

2003; Ernst 2005; McGrath et al. 2006); and the process could be deemed ‘slow’ from a 21st century human perspective (Neidorf 1996).

A global agreement on the use of plants in decontaminating heavy metals is in a state of flux, with many contradictory arguments prevailing currently on the ability of plants to efficiently decontaminate heavy metals in sediments and water. The potential success of phytoremediation is often drawn from unrealistic extrapolations 27

of pot experiments, overzealously interpreting the possibilities of metal extraction

(Vangronsveld et al. 2009). Nonetheless, phytoremediation is worthwhile because the data accumulated so far on the ability of plants to either remove contaminants from the environment or render them incapacitated indicate its potential as a sustainable option (Watanabe 1997), (Table 3). Ways to accelerate remediation by improving plant capabilities through conventional breeding and molecular techniques are currently being explored (Bizily et al. 2003) and will be summarised in this review.

Phytoremediation is a relatively low-operating cost effort compared with other remediation efforts (Glass 2000). For example, the treatment of 3800 L of water contaminated with trinitrotoluene (TNT) and its breakdown products in CWs in

North America cost as follows: activated carbon-adsorption, US $0.40–1.00/L; ultraviolet (UV) oxidation, 1.00–3.00/L; electron-beam destruction, 1.00–5.00/L; whereas phytoremediation would have cost only 0.64/L (Medina & McCutcheon

1996). To reduce soil lead from 1.4 g/kg to 0.4 g/kg in 10 years would have cost $27

000, whereas land-filling method would cost $1 620 000 and soil leaching $790 000

(Wu et al. 2010). These examples offer insights into the cost effectiveness of phytoremediation. Time is, certainly, a limiting factor in phytoremediation because the process is slow (Neidorf 1996). Despite this limitation, the range of environmental benefits, viz. improved biodiversity, soil protection, carbon sequestration, watershed management, diverse sources of energy and aesthetics

(Dickinson et al. 2009), stability, and sustainability, not offered by the other

methods, cannot be gainsaid. Phytoremediation of mine waters is usually deployed in conjunction with CWs at mine sites to remediate accumulating heavy-metal leachate.

CWs are intentionally developed systems that mimic natural sedimentation process, enabling microbial- and plant-based actions, and concurrently enabling a greater level of manipulation than what would occur in a natural wetland (Rai 2008).

Moreover, the success of phytoremediation varies with the plant’s ability to intercept, absorb and accumulate metals, (Ernst 1996), and therefore, each application needs to be tailored considering local strengths, weaknesses and societal requirements. The mechanisms of phytoremediation can be compared with the various physical and chemical alternatives available for management of heavy metals. These have been reported previously, for example Gräfe & Naidu (2008),

Dermont et al. (2008), Pavel & Gavrilescu (2008), Rai (2009, 2012), Vangronsveld et al. (2009) and Juwarkar et al. (2010), and are summarised in Table 2. The objectives of either phytoremediation or other methods are to contain, stabilise or remove the targeted metals. The use of plant-based remediation in wetlands is a sustainable option because it allows continuous capture and immobilisation of metals without the need for electrical pumping and processing systems. Therefore, we suggest that phytoremediation is the only option that offers sustainable immobilisation or extraction of metals from metal-contaminated wetlands. Treatment of heavy metal leachate generated in waste rock dumps at mine sites inevitably requires an in situ solution to treat heavy-metal leachate exiting the dump, as in most cases, relocation of the dump for treatment is not an option. Additionally, the

solution must have the capability to function over the long term, as poorly constructed waste rock dumps can generate leachate for several hundreds of years.

Without preventive measures, the Richmond Mine at the Iron Mountain, California would generate heavy-metal leachate with pH < 1, containing dissolved metals for

3000 years (Nordstrom & Alpers 1999). Modern mining organisations, therefore, invest heavily in reducing the generation of heavy-metal leachate. Additional management efforts launched by mining companies to grade and selectively place the waste rock in dumps thus minimise water percolation (Maddocks 2009), which can reduce – to a large extent – the burden placed on CWs. By placing the rock material that has the potential to generate acid within the core of the dump, leachate generation can be reduced (Hitch et al. 2010). Coordinated efforts by environmental engineers and environmental biologists are enabling successful removal of heavy metals from the leachate, such that ‘clean’ water can be released into surrounding waterways, complying with local environmental regulations. The performance capability of two active CWs at mine sites is presented in Tables 4 and 5.

Table 2. Comparison of conventional and phytoremediation methods for treatment of heavy metals

Medium Remediation method Advantages Disadvantages Cost (US―$/ton) Soil, or Conventional dredged Vitrification Mobility of contaminants reduced Needs intensive energy and high temperatures 400―870 sediment Incineration Contaminant and volume reduced Metals not destroyed and end up in flue gases 250―506 Soil washing Reduces volume of contaminant Contaminant toxicity unchanged, waste stream 60―245 Soil vapour extraction Treatment times (6―48 months) generated –– Electrokinetic Minimal impact on environment Costly treatment of extracted vapour 15―200* (soil removal not required). Metals Effectiveness reduced by alkaline soils removed from soil Wastewater Conventional Chemical precipitation Low capital costs, simple operation Sludge generation and cost of sludge removal and –– Coagulation-flocculation Shorter time to settle out solids disposal –– Dissolved air flotation Low cost, short hydraulic retention time Sludge production and cost of sludge removal and –– Ion exchange No sludge generation disposal –– Ultrafiltration Smaller space requirement Subsequent treatments needed for heavy metal removal –– Nanofiltration Lower pressure than reverse osmosis High capital costs –– Reverse osmosis Able to withstand high pressure High operational cost, prone to membrane fouling –– Costly, prone to membrane fouling High energy consumption Soil, or Phytoremediation dredged Low cost, low energy input, sustainable Seasonal growth, biomass disposal, only suitable for 25―100 sediment, or practice, can be employed in situ, low to moderate contamination, application limited to Wastewater provides beneficial habitat, physical soil rooting depth; rate of decontamination can be slow protection, low maintenance requirement, high social acceptance

*This value is expressed ‘per yd3’.

Table 3. Accumulation of heavy metals by wetland plants in field, glasshouse and hydroponic trials.

Species Plant developmental Trial condition Metal Roots (R), or rhizomes(RH) Shoots(S), or leaves(L), or Reference stage’ flowers (F) Phragmites Seedling Glasshouse or Cu 8883 ± 1010 µg g-1 (R) 445 ± 22 µg g-1 (S) Ye et al. (2003) australis Hydroponic Mature Field Cu 230 ± 12 µg g-1 (R) 59 ± 4.2 µg g-1 (S) 25 ± 12 µg g-1 (Rh) 37 ± 3.2 µg g-1 (L) Phragmites Mature Field Pb 745.13 mg/kg (R) 56.74 mg/kg (S) Ashraf et al. australis 129 mg/kg (L) (2011) 110.68 mg/kg (F) Cu 633.46 mg/kg (R) 345.91 mg/kg (S) 426.78 mg/kg (L) 352.39 mg/kg (F) Zn 356.90 mg/kg (R) 211.40 mg/kg (S) 228.34 (L) 140.30 (F) Phragmites Seedling Glasshouse― Cd 40.7 mg/kg (R) 2.5 mg/kg (S) Stoltz and australis Hydroponic Zn 5608 mg/kg (R) 439 mg/kg (S) Greger (2002)

Cu 485 mg/kg (R) 18.7 mg/kg (S) Pb 1619 mg/kg (R) 28.8 mg/kg (S)

Mature Field Cd 4.6 ± 1.8 mg/kg (R) 1.0 ± 0.5 mg/kg (S) 0.1 ± 0.1 mg/kg (Rh) Zn 1310 ± 901 mg/kg (R) 68 ± 5 mg/kg (S) 80 ± mg/kg (Rh) Cu 80.1 ± 2.8 mg/kg (R) 6.4 ± 5.8 mg/kg (S) 8 ± 2.1 mg/kg (Rh) Pb 523 ± 154 mg/kg (R) 4.1 ± 0.5 mg/kg (S) 8.1 ± 4.5 mg/kg (Rh)

Typha Mature Field Al 350.55 ± 13.71 mg/kg 848.12 mg/kg (S) Hegazy et al. domingensis (R&RH) (2011) Fe 582.44 ± 1.70 mg/kg (R&Rh) 520.38 mg/kg (S) Zn 149.60 ± 0.89 mg/kg (R&Rh) 173.57 mg/kg (S) Pb 20.46 ± 1.33 mg/kg (R&Rh) 27.29 mg/kg (S)

1.5 Heavy metal accumulation and levels of tolerance by plants

For normal growth and metabolism, plants require metals [e.g., Cu, Co, Fe,

Molybdenum (Mo), Mn, Ni, Zn] essentially (Hänsch & Mendel 2009) and a few others

[e.g., As, Cd, mercury (Hg), selenium (Se)] in traces (Rascio & Navari-Izzo 2011).

Plants require Cu to mediate CO2 assimilation; Cu is essential in electron transport mechanisms in photosynthesis and respiration (Demirevska-Kepova et al. 2004).

However, when these metals occur at more than optimal levels as in contaminated land and water sites, plants suffer toxicity resulting in stressed conditions. Stress, including oxidative stress, manifests as reduced vigour and growth inhibition (Levitt 1980)

(Stadtman & Oliver 1991). Excess concentrations of Cu and Cd in the substratum affect germination, seedling growth and lateral-root production in Solanum melongena

(Solanaceae) (Neelima & Reddy 2004). Excess of Mn in leaves reduces photosynthetic efficiency (Kitao et al. 1997). Mn toxicity manifested as necrotic spots on leaves, petioles and stems of Glycine max (Fabaceae) (Wu 1994).

Plants used in metal-inundated sites can either ‘indicate’, ‘exclude’ or ‘accumulate’ metals. Accumulation capability builds on the genetically driven physiological strengths of plants in absorbing and storing metals from the sediment and water

(Knight et al. 1997). Detoxification of metal ions is the principal function. Plants such as Festuca rubra cv. Merlin (Poaceae) are valued for their capability to exclude and stabilize erosion-prone metal-contaminated soils (Wong et al. 1994). Thlaspi caerulescens (Brassicaceae) and Viola calaminaria (Violaceae) are examples of established accumulator species, which do not exclude metals from entering the root, 34

but they detoxify high-metal levels on accumulation in cells (Baker 1981). The excluder species retain the absorbed metals in their roots by altering membrane permeability, changing binding capacity of the tonoplast and cell walls, and chelation of the metal with a ligand, for example citrate and malate (Lasat 2000). The accumulator species transport and retain metals in their shoot tissues proportional to metal levels occurring in the substratum (Ghosh & Singh 2005). Phragmites australis

(Poaceae) – an accumulator species – shows positive linear relationships of Cd, chromium (Cr), Cu, Hg, Mn, Ni, Pb and Zn between those stored in the aerial organs and levels of concentrations in the water and sediment (Bonanno & Giudice 2010), suggesting that P. australis is also an indicator. The hyperaccumulators – plants that can store large quantities of metals – store metals in shoots [up to > 1.0% of Zn and

Mn, up to 0.1% of aluminum (Al), As, Se, Ni, cobalt (Co), Cr, Cu and Pb, and up to

0.01% of Cd] (Baker & Brooks 1989). Hyperaccumulation of metals capability is known in about 450 species of flowering plants (for details, see Sarma 2011), which take up, transport and sequester antimony (Sb), As, Cd, Co, Cu, Mn, Pb, Se, titanium

(Ti), and Zn. Hyperaccumulating plants differ from excluder species by actively taking up and translocating large volumes of either one or more metals from roots to aerial organs, particularly the leaves, at concentrations 100–1000 times greater than those known in non hyperaccumulating, excluding species (Rascio & Navari-Izzo

2011). Plants are useful in bringing both contaminated terrestrial and aquatic systems within regulated limits. Nevertheless, to date, most trials done on metal hyperaccumulating plants have been on nonaquatic plant taxa. Either Thlaspi caerulescens or species of Salix (Salicaceae) grown in contaminated soils reduced Zn

concentrations from 440 mg/g to < 300 mg/g, thus meeting the Zn concentration standards as required by the Commission of the European Community (McGrath et al. 1993; Robinson et al. 1998; Van Nevel et al. 2007).

1.6 Physiological mechanisms in phytoremediation

The success of phytoremediation as a sustainable heavy metal decontamination strategy rests on the plant’s capability in not only absorbing and storing metals in them but also in continuing to grow and produce a fertile progeny. The underpinning physiological principle is their tolerance to metals. Evaluations of the levels of tolerance have enabled breeding metal-tolerant varieties for use in the revegetation of leachate-storage sites (Smith & Bradshaw 1979; Baker 1987). Heavy-metal tolerance in plants is facilitated by either reduced uptake – useful in site stabilisation – or enhanced internal sequestration – useful in phytoextraction. Thlaspi caerulescens,

Nicotiana tabacum (Solanaceae) and Zea mays (Poaceae) (Kayser et al., 2000) possess heavy-metal sequestering capabilities, such as binding the metal ions to their cell walls, reduced efflux pumping of metal ions at the plasma membrane, chelating metal ions in the cytosol with different ligands [e.g., phytochelatins, metallothionein

(MT) and metal-binding proteins] and sequestering metal ions in vacuoles enabled by transporters in the tonoplast (Yang et al. 2005).

Plants absorb metals from the soil matrix using proton pumps in the plasma membrane and ligands chelating the metal (Mukhopadhyay & Maiti 2010) (for illustration of

phytoremediation processes, see Singh et al. 2003; Pilon-Smits 2005; Prasad et al.

2010). Bioavailable metals move through the root symplast crossing the plasma membranes of root endodermal cells. If the metal were to be translocated to aerial tissues, then it enters xylem. To achieve that, metals pass through the endodermal cells enabled by the membrane pump, which supplies energy [adenosine-5′- triphosphate

(ATP)]. Once in xylem, xylem sap flow mechanically transports metals to leaves passing through cell membranes. The metal is either sequestered in one of the subcellular compartments (e.g., cell wall, cytosol, and vacuole) or volatilised through the stomata (Davis et al. 1998). For example, Hg is volatised through conversion to the elemental form in transgenic Arabidopsis (Brassicaceae) and Liriodendron tuliperfera

(Magnoliaceae) that encode the gene for mercuric reductase (MerA) (Rugh et al. 1998).

Se has been shown to be volatilised following conversion to dimethyl selenide (DMSe) by rhizosphere bacteria and cyanobacteria (Neumann et al. 2003). Such processes liberate not only essential metals but also toxic metals from soils (Lasat 2000).

Plants growing in heavy-metal-contaminated sites display either sensitivity – resulting in either injury and death, or resistance – displaying defence reactions. The latter enables survival and reproduction (Levitt 1980). Equipped with appropriate subcellular mechanisms, plants detoxify metals and overcome the heavy-metal- induced stress. Cardaminopsis halleri (Brassicaceae), a metal-tolerant taxon raised on

Zn and Cu-contaminated soil, includes electron-dense metallic precipitates of Zn, Cu,

Sn, Fe and Al on leaf surfaces, of Zn, Cu and Sn in intercellular spaces, and Zn, traces of Cu and Fe on the walls of xylem elements (Macnair & Smirnoff 1999). Vacuoles

are the principal storage compartments of metals in the metal-tolerant plants. The cytoplasm and nuclei of leaf cells of C. halleri include precipitates of Zn and silica as

Zn silicates. In the mesophyll cytoplasm of Cardaminopsis, Zn gets transiently accumulated as different silicates. Zn gets translocated to vacuoles and stored in an unknown form (Neumann & zur Nieden 2001). A second silica and Zn uptake mechanism beyond the symplastic and apoplastic movements is also known in C. halleri, where the pinocytotic vesicles along the plasma membrane and tonoplast enable a direct translocation of silica and Zn from extracellular compartments into the vacuole. Formation of Zn silicates is a part of tolerance mechanism and could be responsible for nullifying the effect of Zn toxicity in Cardaminopsis (Neumann & Zur

Nieden 2001).

Mycorrhizae are gaining relevance as active participants in heavy-metal tolerance in plants and in phytoremediation. Cd-tolerant isolates of Glomus mosseae (Glomeraceae) influence the uptake, transport and immobilisation of Cd in roots of Trifolium subterraneum (Fabaceae) (Joner & Leyval 1997). Mycorrhizal association improves tolerance of trace metals in contaminated soils growing Glycine max and Lens culinaris

(Fabaceae) (Jamal et al. 2002). The supposition here is that the mycorrhizae enhance the ability of the plant to cope with water stress induced by either nutrient deficiency or drought (Schreiner et al. 1997). Chitin and cellulose –mycorrhizal wall materials – offer the critical potential for metal binding (Tsezos et al. 1997). In many instances, extracellular polymeric substances such as exopolysaccharides that arise as vesicles on

the cell membrane have been shown to facilitate metal immobilisation (Cao et al.

2011).

1.7 Genetic engineering for enhanced heavy-metal storage in accumulating and tolerating plants

Tolerance to pollution, fast-growing, competitive, hardy and producing high biomass are desirable properties of plants for use in phytoremediation (Pilon-Smits 2005).

With advancing knowledge of molecular physiology of heavy-metal accumulating and -tolerating plants, development of genetically engineered plants enabled with greater capabilities to store heavy metals has ushered in a new era in phytoremediation in the early 2000s. Transgenic plant taxa successfully extract Cd,

Pb, Cu, As and Se from soil and water, and store them in their aerial organs.

Transgenic plants use metal transporters and generate a suite of enzymes to metabolise metal-detoxifying chelators, such as MTs and phytochelatins. For example, those plants that can synthesise MerA and organomercurial lyase (MerB), convert the organo-Hg to metallic Hg to be volatised through leaf surface.

Volatilisation of Se compounds was promoted in plants overexpressing the gene- encoding enzymes involved in production of DMSe (Kotrba et al. 2009). Because glutathione is the precursor for different heavy-metal-binding peptides and also because we knew that glutathione has the ability to prevent cell damage caused by reactive oxygen species (ROS), which is critical in plant defence against a range of abiotic threats, its metabolism is relevant in understanding tolerance and sequestration of heavy-metal ions. Individuals of Brassica juncea (Brassicaceae), overexpressing 39

the selenocysteine methyltransferase (SMT) gene from the Se-hyperaccumulator

Astragalus bisulcatus (Fabaceae), have been shown to have a greater capacity to store

‘methionineseleno- cysteine’ complex and tolerance to selenite and related Se compounds (LeDuc et al. 2004, 2006). The role of glutathione and its derivatives – different phytochelatins – in heavy-metal tolerance have been illustrated in Cd- sensitive mutants of Arabidopsis (Cobbett 2000). Using genetically engineered B. juncea with an overexpressed Escherichia coli gshII gene encoding glutathione synthetase (GS) shows that the engineered B. juncea accumulated more Cd than the unengineered wild taxa (Zhu et al. 1999).

Genetically engineered plants show greater levels of tolerance to Cd at both the seedling and mature plant stages. Most importantly, Cd accumulation and tolerance correlated with expression level of the nucleotide sequence for gshII. The MT-coding gene (PsMTA) from Pisum sativum (Fabaceae) was transposed into Arabidopsis thaliana, and the transformed plants accumulated several-fold higher values of Cu

(Evans et al. 1992). A type-2 metallotheonein-coding gene (tyMT) from Typha latifolia (Typhaceae) was transferred to A. thaliana, and the transgenic A. thaliana showed a greater level of tolerance to Cu and Cd (Zhang et al. 2004). Transgenic plants that express MTs have been scored for enhanced tolerance, accumulation and distribution of Cd. A human MT-II gene was transposed into seedlings of N. tabacum and B. napus; growth of these remained unaffected up to 100 mmol/L of Cd concentrations (Misra & Gedamu 1989). An increased level of Cd tolerance – up to

200 mmol/L – is known (Liu et al. 2000), and an altered distribution of Cd has been

observed in transgenic plants that express MTs. For example, the human MT-II gene and the ‘MT-II fused with b-glucuronidase gene’ were expressed in N. tabacum under control of the cauliflower mosaic virus (CaMV) 35S promoter with a double enhancer

(35S2) (Mejáre & Bülow 2001). In vitro grown transgenic seedlings expressing the fusion protein accumulated 60–70% less Cd in their shoots than the control plants did

(Elmayan & Tepfer 1994). Transferring genetic material from low-biomass hyperaccumulating species into high-biomass species is also being currently explored

(Pilon-Smits 2005). For example, genetic engineering efforts at United States

Department of Agriculture (USDA) (Beltsville, MD, USA) are aimed at improving the biomass production of Zn-tolerant Thlaspi caerulescens (Fulekar et al. 2009). The use of genetically engineered plants for clean-up of contaminated sites is however restricted by legislation and concerns of the community. The biosafety risk of horizontal gene transfer by transgenic species of Populus were assessed through field trials of Cu accumulation in mining areas of Saxonia-Anhalt (Germany); results indicated that the transgenic species of Populus are genetically stable with no indications of any impact on the environment to date (Peuke & Rennenberg 2005).

1.8. Efficacy of wetland plants in remediating heavy-metal-containing wastewater

The major environmental problem facing the mining industry is with the oxidation of pyrites (e.g., FeS2), which leads to generation of metalliferous leachate (Egiebor &

Oni 2007). Problems involving mine leachate occur due to most base metal, precious metal, uranium and coal deposits containing pyrites either in the ore or in the surrounding waste rock (Jensen & Bateman 1979). Pyrite oxidation is complex

involving chemical, biological and electrochemical reactions, and varies with environmental conditions. For instance, the rate of oxidation is determined by pH,

PO2, specific surface structure of pyrite, presence and/or absence of bacteria (e.g.,

Ferrobacillus ferrooxidans) and clay minerals, and hydrological factors (Evangelou

& Zhang 1995). Leachate consists of varying levels and combinations of dissolved metals (Mayes et al. 2009), acid rock drainage (ARD, pH < 6), neutral mine drainage

(NMD) and saline drainage (SD, pH > 6) (Plumlee et al. 1999) (Figure 2). S2-based minerals, such as pyrite (e.g., FeS2), that usually occur associated with the waste rock enable release of heavy metals in solution (Salomons 1995).

Mine sites generating heavy-metal leachate require an effective and sustainable means of treating heavy metal leachate. In many historical mine sites, however, leachate has been allowed to exit mine sites with no treatment, leading to widespread environmental destruction. The Rum Jungle uranium–Cu project in Northern

Australia (13°02′47″S:131°01′39″E) commenced in the 1950s, continues to be an environmental threat to neighboring waterways (Mudd & Patterson 2010). The Rum

Jungle project was operated on a ‘production’ basis, and environmental impacts were not recognised at that point of time (Lichaz & Myers 1977). Despite the scale of the environmental threat because of the release of approximately 1 mL/day of acidic liquid effluent (pH 1.5), and metal and radionuclides being identified in the 1950s, no remediation attempts were made until the 1970s (Davy 1975). The Australian government commenced remediation between 1982 and 1986, budgeting $18.6 million (Richards et al. 1996). Despite these efforts, the site still produces annual

2- pollutant loads at 4–12 t Cu, 3–7 t Zn and 1250–4800 t sulphates (SO4 ) (Mudd &

Patterson 2010).

Treatment of mine leachate using wetlands was first identified in the Appalachian coal mines in the 1980s. Drainage from these coal mines was amenable to treatment by natural wetlands because of a combination of small flows and diverse chemical composition (Sobolewski 1999). The use of CWs for treating mine wastes are a viable alternative to chemical dosing, sulfidogenic bioreactors, limestone drains and permeable reactive barriers (Johnson & Hallberg 2005). Several examples of CWs for metal removal have been reported at active mine sites (Brooks 1991; Dunbabin &

Bowmer 1992; Ryan & Hosking 1992; Eger 1994; O’Sullivan et al. 2003). While vegetation is a vital component in removing heavy metal from mine leachate, the majority of contaminants accumulate within the substrate (Wood & Shelley 1999).

Heavy metals occur in contaminated wetlands as either suspended matter in the water column, bound to particulate matter accumulated in the sediment (Dunbabin &

Bowmer 1992) or stored within plant matter. In solution, metal ions can form complexes and share one or more pairs of electrons with ligands or chelating agents, and form a ligand-metal complex (Boye & Van Den Berg 2000). Humic and fulvic acids are the most commonly occurring ligands in natural waters (Pagenkopf 1978).

Metals have a high affinity for humic acids, organo clays and oxides coated with organic matter; the chemical forms of metals in soils and sediments are: water-soluble free ions, for example Zn2+, inorganic complexes and organic complexes; exchangeable metals bound to soil surfaces by cation exchange process but may be

readily displaced due to the water-soluble form; precipitated as inorganic compounds,

2- for example metal oxides, hydroxides(-OH) and carbonates (CO3 ); complexed with large molecular weight humic materials (Gambrell 1994); adsorbed or occluded, for example Mn and Al, and most trace metals are adsorbed or occluded by Fe oxides

(FeO, Fe2O3); precipitated as insoluble sulphide reacting with trace and toxic divalent metals forming insoluble metal sulphide precipitates; bound within the crystalline lattice structure of clay minerals by isomorphous substitution, most of these metals are unavailable and only become available after mineral weathering (Gambrell 1994).

The anaerobic and/or reducing conditions in wetland soils tend to converge towards pH 7 regardless if the underlying soil conditions were acid or alkaline prior to flooding; these near neutral conditions favour metal immobilisation (Ponnamperuma

1972).

CWs in mine sites can be designed to function as long-term storage sinks (Dunbabin

& Bowmer 1992) slowing the distribution of heavy metals into surrounding waterways. In this context, the metal may still pose a threat but is immobilised and stored within the wetland by biogeochemical processes. Further manipulation can be achieved depending on plant selection and traits; the aim should be to focus on metal immobilisation (phytostabilisation) or removal (phytoextraction). Knowledge of the long-term fate of contaminants in wetlands is however limited; data are available of the removal efficiency in the short-term and long-term performance, and storage ability is limited (Athay et al. 2003) because of many systems only being in operation for between 10 and 15 years (O’Sullivan et al. 2004). The Compañia Minera

Antamina S.A. (CMA) mine in the Tucush Valley in Central Andes, Peru

(9°45′57″S:77°06′00″E) CWs for filtering NH3, N03, Cu, Mo and Zn leachate exiting waste rock dumps in 2005–2006. The created wetland (6 ha) comprised sediment ponds, serpentine channels and wetlands designed to treat up to 115 L/s of leachate.

Throughout the trial period (2006–2009), the water parameters measured at the outlet complied favourably with water quality requirements dictated by the Ministerio de

Energía y Minas del Perú (Table 4) (Strachotta et al. 2009). Two pilot-scale wetlands were constructed for monitoring at an active Pb-Zn mine (the Tara Mines) Navan,

Ireland (53°39′12″N:6°42′50″W) in 1997 to remove different metals and S04-s from wastewater. The CWs at Tara Mines removed 81% S04-s, 32% Pb and 74% Zn

(Table 5) (O’Sullivan et al. 2004). CWs have also been used to remove heavy metals from industrial wastewaters (Khan et al. 2009; Galletti et al. 2010). CMA mine, Peru is an example of remediation measure established while mining activities occur, a term we refer to as ‘remediation concurrent with waste generation’ enabling a more effective and less burdened remediation system. The mine closure process should be seen as a component of a complete life cycle obligation and factored into mine activities from the start until closure (Lambeck 2009).

Figure 1. Ficklin diagram showing acid-rock drainage, neutral mine drainage, saline drainage (Plumlee et al. 1999).

Table 4. Wetland treatment of mine drainage from CMA mine in Tucush Valley waste dump (Estimated parameter concentration reduction - based on average loading rates) (Strachotta et al. 2009).

Metal Inlet average loading Guideline* Outlet average loading Reduction % Mo 0.23 kg/day 0.2 mg/L 0.08 kg/day 65 Zn 0.27 kg/day 0.215 mg/L 0.13 kg/day 51 * Ministerio de Energía y Minas del Perú water quality discharge parameters.

Metal Inlet Outlet (mean) Reduction % Range Reduction % 2- -1 -2 -2 SO 4 900 mg L 10.4 g day 31 0―42 g m day 81 Pb 0.15 mg L-1 1.9 mg -2 day 32 0―6.6 mg m-2 day 0―64 Zn 2.0 mg L-1 18.2 mg m-2 day 74 0―70 mg m-2 day 0―99

1.8.1 Wetland species selection for phytoremediation

Wetlands are characterised by anaerobic conditions because of the reduced diffusivity of gases in water; gas exchange between soil and the atmosphere is approximately 10 000 times slower than that in air (Armstrong 1978). Once flooded, oxygen is rapidly consumed by the metabolism of microbes and chemical oxidation (Batty 2003). Oxygen deficiency and phytotoxin accumulation because

2+ 2+ of the creation of Mn , Fe and H2S by anaerobic microbes are major characteristics of water-logged soils (Armstrong et al. 1991). The ability to survive in these environments is dependent on the ability to move oxygen from shoots and leaves to roots and rhizomes. Plants also have several adaptations to enable survival when shoots are submerged during floods. When shoots are completely submerged and O2 is exhausted, anoxia can occur; rhizomes, tubers and shoots of some wetland species tolerate these conditions by forming anoxic cores deep within their tissue surrounded by zones of hypoxic or aerobic tissue, which derive O2 by diffusion from adjacent water-logged soil or airspaces (aerenchyma) within ground tissues (Atwell et al. 1999). Wetland plants have evolved to cope with such conditions; however, the conditions imposed by excessive metal concentrations create conditions outside of the natural environmental conditions characteristic of natural wetlands (Otte 2001). Dryland plants are unable to tolerate these conditions, that is, water-logging and low availability of oxygen to roots (Otte 2001). In wetland plants, internal gas spaces provide a conduit (convective flow) for air to move down to the root zone through gas convection (Armstrong et al. 1992).

Although wherever there is internal O2 transport, there will be a potential for radial

O2 loss (ROL) from root to soil, the oxidative removal of soil-borne phytotoxins is an additional survival factor (Armstrong 1982). ROL oxygenates the rhizosphere and has the capacity to increase sediment Eh (Wright & Otte 1999), and Chen &

Barko (1988) have found an increase in soluble Fe consistent with higher Eh, while

Doyle & Otte (1989) show that ROL could decrease metal mobility. Furthermore,

ROL generation has been found to be dependent on each individual species, for example Glyceria fluitans has little effect on sediment chemistry because of limited

ROL (Wright & Otte 1999). It is noted that a consensus on the potential of ROL to increase metal mobility through an influence on the increase or decrease on soil Eh and pH is yet to be reached.

To date, no particular aquatic species has been nominated capable of storing sufficient quantities of a metal to be ranked as a hyperaccumulator (Marchand et al.

2010), although Typha species in particular are known for their heavy-metal tolerance and high biomass production (Pilon-Smits 2005). Wetland plants, in general, have a tendency to store metals below ground and limit their translocation to shoots (Batty & Younger 2004). A good accumulator would be rated on its ability to take up > 0.5% dry mass (DM) of a given element and establish a bioconcentration factor (BCF) in excess of 1000 (Zayed et al. 1998).

Hyperaccumulator species are useful in phytoextraction, wherein metals are accumulated by a plant species known to transport and concentrate heavy metals into harvestable, high–biomass, above-ground shoots (Lasat 2002). Although, those

working to develop long-term, low-to-nil maintenance wetlands may prioritise metal storage within underground root tissues and therefore value the quantity of root accumulation. Plant material for CWs at mine sites is usually sourced from either nursery stock or seeds obtained from nearby natural wetlands. Selection of wetland plants is best done from local, endemic wetland species. Taxa naturally colonising waterways within and around mine sites are often planted in CWs at mine sites and include species of Typha (Typhaceae), Juncus (Juncaceae) and

Phragmites (Poaceae), and Cyperus, Schoenoplectus, Restio, Eleocharis and

Fimbristylis (Cyperaceae) (Dunbabin & Bowmer 1992). Metal tolerance evolves in plants after exposure to increased metal concentrations (Antonovics et al. 1971;

Macnair & Baker 1994). It is suggested that this is not the case with wetland plants and that they may be innately tolerant of high levels of metal concentrations (Otte

2001). To establish if the wetland plant Glyceria fluitans developed Zn-tolerant ecotypes, populations were collected from high Zn concentrations (253 mmol/g) in mine tailings and low Zn concentrations (2.4 mmol/g) from a natural waterway; both populations were grown in mine tailings, and there was no difference between growth, biomass and metal uptake, suggesting that both populations are equally tolerant despite their previous growing conditions (Otte 2001). However, when exposed to excessive metal concentrations, wetland plants display limitations.

When treated with increasing concentrations of Cu at 400 µg/g, biomass and root- growth inhibition occurred in Avicennia marina (Verbenacae), and total inhibition of seedling emergence occurred at 800 µg/g (MacFarlane & Burchett 2002). Not only do plants in CWs remove heavy metals from the sediment and wastewater but

they also catalyse biogeochemical processes by providing organic matter through senescence and organic compounds through exudation (Jenssen et al. 1993).

Accumulated organic matter of plant source in wetlands adsorbs metals, and the

‘metalorganic matter’ complexes function as a carbon source for bacterial metabolism (Beining & Otte 1996; Jacob & Otte 2003). Ionic forms of heavy metals, for example Cu and Zn, also form complexes with organic matter. Because plants cannot absorb large metal complexes, such complexes are unavailable to plants (Reichamn 2002).

1.8.2 Effects of heavy metals on plants

Excessive levels of heavy metals induce plant stress, resulting in either reduced vigour or total inhibition of plant growth (Levitt 1980). Certain plant species exhibit sensitivity to the effects of heavy metals, which results in either death or injury, while those capable of tolerating such conditions express resistance, surviving and producing tolerant offspring (Baker 1987). Plants require Cu for CO2 assimilation, this element is also an essential component of proteins like plastocyanin of photosynthetic system and cytochrome oxidase of respiratory electron-transport chain (Demirevska-Kepova et al. 2004). However, excessive exposure to Cu generates oxidative stress and reactive oxygen species (ROS)

(Stadtman and Oliver 1991). Copper and Cd in combination affect germination, seedling growth and lateral root production in Solanum melongena adversely

(Neelima and Reddy 2004). Accumulation of excessive Mn in leaves reduces photosynthetic rate (Kitao et al. 1997). Zinc and Cd decrease growth and 51

development, metabolism and induce oxidative stresses in Phaseolus vulgaris

(Cakmak and Marshner 1993), and Brassica juncea (Prasad 1999).

1.9 Metal removal processes in wetlands

Heavy-metal removal mechanisms in wetlands occur via three main processes: (1) soil and substrate, (2) hydrology, (3) vegetation. Either permanent or periodic saturation results in anaerobic conditions in the soil under which wetland biogeochemical processes occur and regulate development of wetland substratum, which support a plant community adapted to living in soils saturated with water

(Mitsch and Gosselink 1993). Removal of heavy-metal ions in wetlands occurs by the interacting processes of settling, sedimentation, sorption, co-precipitation, cation exchange, phytoaccumulation, microbial activity and plant uptake.

1.9.1 Wetland plant removal of metals

Elemental uptake by wetland plants was initially thought to occur solely through leaves whereas the roots acted as anchors (Jackson 1988). Studies made in the

1970s and 1980s demonstrated that macrophytes accumulated heavy metals relative to concentrations in the surrounding water (Kovacs 1978; Heisey and Damman

1982; Miller et al. 1983). Concentrations of metals accumulated in plants relative to water concentrations led to the understanding that plants accumulated metals from water (Dietz 1973; Kovacs 1978; Kovacs et al. 1973). However, the current understanding is that the source of most element nutrition for rooted macrophytes is 52

removed from sediments which house root systems (Jackson 1998). Following root uptake elements are transported to above-sediment tissues (Denny 1972).

1.9.2 Physical removal of metals

Settling and sedimentation are the principal processes in the removal of heavy metals associated with particulate matter in acid-mine water (Horner 1995). Heavy metals are removed from waste water and trapped within the wetland sediments

(Hares and Ward 2004). Efficiency of removal of suspended solids is proportional to the particle-settling velocity and the physical length of the wetland (Johnston

1993). For particles less dense than water, sedimentation occurs after flocculation, which enables them to settle more rapidly than individual particles; flocculation is accelerated by high pH, concentration of suspended matters, ionic strength and high-algal concentration (Sholkovitz 1978; Matagi et al. 1998).

1.9.3 Chemical removal of metals

A range of chemical processes is involved in the removal of heavy metals from wetlands. Metal ions can be bound in either short-term retention or long-term immobilisation through chemical adsorption as metal ions are transferred from water to the soil, i.e., from the solution phase to the solid phase (Sheoran and

Sheoran 2006). Heavy metals can also be adsorbed to clay and organic matter electrostatically, because cation exchange involves physical attachment of cations to clay and organic matter (Balasoiu et al. 2001). In excess of 50% of heavy metals 53

can be readily adsorbed onto particulate matter in wetlands and therefore removed from water by sedimentation (Muller 1998). Lead and Cu are strongly adsorbed, whereas Zn, Ni, and Cd are likely to be more labile and rendered bio–available

(Alloway 1990). Once adsorbed on to either humic or clay colloids, heavy metals remain as atoms, although their speciation may change with time with changes in the sediment conditions (Weibner et al. 2005). Iron, Al, and Mn form insoluble compounds through redox reaction, forming a variety of oxides, oxyhydroxides, and hydroxides (Woulds and Ngwenya 2004). Manganese removal is difficult due to oxidation occurring at a pH close to 8 (Stumm and Morgan 1981). Formation of

2+ insoluble heavy metal precipitates, e. g., Mn oxidised and precipitated as MnO2, can limit the bioavailability of heavy metals (Sheoran and Sheoran 2006).

Although, the precipitated heavy metals can be released back into solution to restore equilibrium, metals incorporated into mineral lattices are in principle unavailable to biota (Dunbabin and Bowmer 1992). Copper, Ni, Zn, and Mn can co-precipitate with secondary minerals such as FeO2 and Co; Fe, NI, and Zn co- precipitate in MnO2 (Stumm and Morgan 1981; Noller et al. 1994). Alkaline conditions are necessary for co-precipitation of cationic metals such as Cu, Zn, Ni, and Cd, with FeO2 and MnO2 by co-precipitation and adsorption (Drever 1988).

2.0 Cadia Valley Operations

Cadia Valley Operations (CVO) (Figure 1) occurs at 25 km south-west of Orange in the Central Tablelands of New South Wales. The mining lease area covers

approximately 3,100 ha in the Cadia Valley, which drains the southern portion of

Mount Canobolas, located 13 km to the north. The land surrounding the mining lease is used for cattle and sheep grazing for meat and wool, with dairying, horse breeding, mixed crops, and apiculture occurring on the undulating, cleared, and grassy landscapes.

Figure 2. Aerial image of Cadia Valley Operations, Orange, NSW (Scale: 1 cm = 0.5 km).

The mining lease belongs to Newcrest Mining Limited (Newcrest, hereafter).

Mining operations at CVO commenced in August 1998 following the discovery of low-grade Au ore in 1992. Low-grade Au and Cu ore is excavated from Cadia Hill 55

open pit and high-grade Au and Cu ore from an underground mine (Ridgeway).

Cadia Valley Operation also has a new underground mine ‘Cadia East’ within the boundary of the current Newcrest mining lease. The locations of the underground mining lifts are north of the south waste-rock dump (SWRD) and the east of the current Cadia Hill open pit, the project commenced in 2013. The SWRD will incorporate approximately 11.4 mt of waste rock from the Cadia East expansion in the current footprint and a further 2 million tonne from the Cadia Hill open pit and

Ridgeway underground mine (OKC 2010). Conventional truck and hydraulic excavators are utilised to extract blasted ore from the pit. The Ridgeway mine uses sub-level cave-mining method, wherein the ore body is mined from the top, employing drilling and blasting. The ore-processing plant comprises a crushing, grinding, and a flotation circuit to recover the gold and copper and produce a gold rich copper concentrate.

2.1 Waste rock dumps at Cadia Valley Operations

Soil and rock excavation from the open pit and underground mining activities at

Cadia Valley Operation (CVO) generates large waste-rock dumps that are located on-site. The excavated soil and rock material is categorised as either suitable for processing; or unsuitable and is left to remain in-situ, or excavated and dumped as waste (Osanloo et al. 2008). High-operating costs can deter mining and processing of the ore, although at times of higher metal prices these wastes may be treated as ore (Hitch et al. 2010). Cadia Valley Operations has two waste-rock dumps, the north waste-rock dump (NWRD) and the SWRD. The NWRD is located to the 56

north of the open-pit and spreads over 45 ha, the SWRD to the south of the open-pit and spreads over 300 ha. The present study specifically focuses on the leachate drained from the SWRD, being the larger of the two; it generates the greatest quantity of heavy metal leachate and represents the most significant potential environmental threat. Approximately 430 million tonne of waste rock has been mined and stored in the waste-rock dumps.

The SWRD occurs in a valley to the south of the Tertiary basalt ridge with the

Cadia Hill open pit to the north-west and the Tailings storage Facilities (TSFs) to the east. Prior to placement of the SWRD the ground surface consisted of two merging creeks and associated valleys. Two leachate ponds, northern leachate pond

(NLP) and southern leachate pond (SLP) are located on the western side of the

SWRD to collect and recirculate leachate exiting the SWRD. Cadiangullong Creek occurs approximately 1000 m west of the SWRD, flowing north---south. Silurian sediments occur under the majority of the SWRD with a small portion in the south west corner of the dump underlain by massive basalt (Tertiary Basalt) fractured at the surface. A fault dipping to the east separates the Silurian sediments underlying the SWRD from the basalt. Surfical clays overlie the Silurian sediments between the Cadiangullong Creek and the SWRD (OKC 2010).

Above ground waste-rock dumps, such as those at CVO remain unsaturated, containing 5–10% water. Any unsaturated zones of sulphidic waste rock are susceptible to acid-rock drainage (ARD) may either seep from the toe of the waste-

rock pile or migrate beneath the pile into groundwater. The activity of each waste- rock dump is time dependant and depends on the physical properties of materials such as porosity, grain size (implying, surface area), diffusion coefficient, gas permeability, hydraulic conductivity and thermal conductivity. The CVO mineralisation is a low-sulphide system, reflected by less than 5% pyrite; the sulphide minerals occurs in two associations, as inclusions in quartz veins and as

‘fine–medium’ ground dissemination in the monzonite porphyry and volcanic host rocks (Johnston et al. 1997). Several measures have been introduced to minimise the subsequent generation of AMD from the waste-rock dumps. Waste rock material is partitioned within the SWRD such that non-acid forming (NAF) waste rock (blue material) is confined to the western portion of the dump. Potentially acid forming (PAF) waste rock (pink material) is confined to the eastern portion of the dump with mineralised waste (yellow material) and low grade ore occur in smaller patches along on the northern areas of the dump (OKC 2010).

2.2 Aims and objectives

The objective of all mining activity is to extract valuable mineral resources.

Throughout the mining process however, unwanted rock material – due to minimal ore content must also be excavated. This waste–rock is stored as dumps at mine sites and contains the sulfide mineral pyrite (FeS2), once exposed to the atmosphere and water pyrite produces acid-mine drainage (AMD). Saline mine-leachate is also generated from pyrite, although the presence of acid-neutralising carbonates within

the waste–rock increases the pH of the leachate. Saline mine-leachate is usually dominated by excess concentrations of Ca, Mg, and S.

Phytoremediation of mine leachate in natural and artificial wetland systems is a proven natural and sustainable means to capture and remove excess concentrations of elements from the mine leachate. Volunteer wetlands surrounding mine sites were discovered in the 1980s in the Appalachian Mountains of United States, and this lead to research into the suitability of plant-based remediation of mine leachate.

Alternative chemical and engineering options are available for treatment of mine leachate, although these methods are not considered a sustainable practice.

Research efforts into phytoremediation of mine leachate using volunteer and artificial wetlands have largely focussed on AMD, whereas phytoremediation of saline-mine leachate has received less attention. Furthermore, limited research has been conducted on the phytoremediation capacity of Australian native wetland plants. A key limitation of phytoremediation of mine leachate is the plants ability to tolerate and accumulate the excessive concentrations of elements in soil and sediment. Excessive elemental concentrations within wetland soil and sediment can inhibit growth and performance of plants and compromise the success of phytoremediation.

Therefore, the aim of this PhD study was to investigate the suitability of several

Australian native wetland plants (Typha domingensis, Baumea articulata, Carex appressa, Phragmites australis, and Schoenoplectus validus) for use in

phytoremediation of the saline-mine leachate. A field site at Newcrest’s ‘Cadia

Valley Operations (CVO)’ Au and Cu mine was selected to conduct field trials, and for collection of plant material and soil and sediment impacted by saline leachate runoff from stockpiled waste-rock dumps. Hydroponic and glasshouse trials were conducted at the horticulture research centre at Charles Sturt University, Orange,

NSW.

The use of plant-based strategies to remediate contaminated dryland and wetland soils and sediment makes use of the naturally occurring processes by which plants and their microbial rhizosphere flora degrade and sequester organic and inorganic pollutants (Pilon-Smits 2005). Phytoremediation has gained acceptance as a cost- effective, non-invasive and sustainable treatment technology over the past 20 years.

Chapter two covers a review of the treatment of heavy metals at mine sites, the review covers plant performance in metal impacted wetlands and present the argument that phytoremediation is the only sustainable treatment option of heavy metal leachate (Chapter 2).

Treatment of mine-waste water was found to occur naturally in wetlands surrounding mines in the Appalachian Mountains (U.S.), this led to experimental work in determining the role wetland plants play in treating mine leachate in the

1980s. Natural colonisation by plants at mine sites in submerged tailings and leachate ponds is a good indication of a given wetland plants ability to tolerate excess concentrations of contaminants occurring in the leachate and sediment. The

wetland plant Typha domingensis has naturally colonised the waterways and leachate ponds at CVO. The aim of chapter three therefore, was to examine the ability of T domingensis to accumulate Cu, Mn, and Zn, and its role in removing concentrations of these metals from leachate as it exited the base of the waste-rock dumps and flowed through stands of T. domingensis before settling in the leachate ponds (Chapter 3).

In addition to the assessment of the natural coloniser T. domingensis, Phragmites australis, Carex appressa, Baumea articulata, and Schoenoplectus validus were also selected. The aim of chapter four was to therefore screen the selected species with a hydroponic experiment to determine their ability to accumulate Ca, Cu, Mg,

Mn, S, and Zn (Chapter 4). The use of hydroponic experiments to screen species for phytoremediation is a cost-effective alternative to field experiments (Migeon et al. 2012). Following the successful hydroponic screening of P. australis, C. appressa, B. articulata, and S. validus, the aim of chapter five therefore was to conduct glasshouse trials with the selected species to determine their photosynthetic performance, as an indication of tolerance to excess concentrations of Ca, Cu, Mg, Mn, Na, S, and Zn (Chapter 5).

The ability of wetland plants to tolerate excess metals and salts is thought to be an innate trait (McCabe et al. 2000). This is due to the natural biogeochemistry of the rhizosphere (Matthews et al. 2005). Wetland soils are reduced and anaerobic, although the rhizosphere is aerobic and oxidised which can facilitate mobilisation

of metals and lead to enrichment of metals immediately surrounding the roots due to movement of air from shoots to roots (Doyle and Otte (1997). Additionally, restricting upward movement of metals and salts from roots to shoots and thus prevent damage to the photosynthetic mechanism is a valuable trait. Therefore, the aim of chapter six was to conduct an assessment of mineral accumulation and translocation traits of P. australis, C. appressa, B. articulata, and S. validus

(Chapter 6).

Chapter seven continues the work of chapter 6 with an assessment of the influence of the selected plants on porewater concentrations of Ca, Cu, Mg, Mn, Na, S, and

Zn in wetland soil and sediment containing excess concentrations of these metals and salts (Chapter 7). Wetland plants differ greatly from dryland plants by being able to survive in saturated soil, phytotoxic substances including Fe2+, Mn2+, and sulfide (Armstrong & Armstrong 1988). In order to tolerate these conditions oxygen (termed radial oxygen loss ROL) transported to roots and not used in respiration is diffused to the rhizosphere and forms an oxidative layer around the root surfaces and prevents plant absorption of reduced Fe, Mn, and sulfide

(Christensen et al. 1994).

2.1 Introduction

Extraction and purification of metal ore at mine sites generates large volumes of heavy-metal wastes, which are usually stored in tailings dams and waste-rock dumps. Such waste-rock dumps generate and release leachate. Leachate consists of varying levels and combinations of dissolved metals and salts (Mayes et al. 2009), acid-rock drainage (ARD, <6 pH), neutral-mine drainage (NMD), and saline drainage (SD, pH >6) (Plumlee et al. 1999). S2– based minerals, such as pyrite

(e.g., FeS2) that usually occur associated with the waste-rock enable release of heavy metals in solution (Salomons 1995). Oxidation of FeS2 involves chemical, biological, and electrochemical reactions, but varies with environmental conditions; the rate of oxidation is determined by pH, PO2 levels, specific-surface structure of

FeS2, presence and/or absence of bacteria such as Thiobacillus ferrooxidans and T. thiooxidans (Acidithiobacillales: Acidithiobacillaceae), clay material, and different hydrological factors (Evangelou & Zhang 1995). The leachate contains dissolved levels of constituents solubilised from the rock, depending on the chemical properties of the waste material and the chemistry of interstitial water (Halford

1999). High levels of Cd, Cu, Cu, Pb, and Fe, and their salts occurring in the leachate can be toxic to organisms (Guilizzoni 1991). Removal of heavy metals

from mine wastewater can be achieved by physical and chemical methods, such as: oxidation, reduction, precipitation, membrane filtration, ion exchange, electrochemical operation, biological treatment, and adsorption (Corami et al.

2007). However, these methods bear several environmental disadvantages: exhaustion of natural resources and energy, generation of large volumes of waste, and emission of CO2 and diesel fumes, while the use of plant-based treatment of mine water is a cheap and sustainable alternative to these measures (Rai 2009).

Treatment of mine leachate using wetlands was first identified in the Appalachian coal mines in the 1980s (Weider & Lang 1982; Bastian & Hammer 1993).

Drainage from these coal mines was amenable to treatment by natural wetlands due to a combination of small flows and diverse chemical composition (Sobolewski

1999). To satisfy state and federal legislative requirements mine operators are required to commit to high-quality rehabilitation of mine-sites upon closure. As such, the use of constructed wetlands for treatment of leachate is an appealing option for mining operations committed to long-term rehabilitation of their mine- sites (Connick et al. 2010). Moreover, mining companies are moving toward towards an integrated treatment approach of managing waste through minimisation and recycling (Dudeney et al. 2012). Constructed and natural wetlands in such contexts function as sinks, removing metals from the solution and retaining them in sediments in a stable, i.e., biologically unusable, form. Several examples of successful treatment of acidic- and alkaline- mine leachate in constructed wetlands are available (Sasaki et al. 2003; An et al. 2011). In a study of a 90-y old abandoned mine-tailings pond in Wicklow Mountains National Park, Ireland

(53°00’59.52’N; 6°22’02.32’W) that was naturally colonised by Glyceria fluitans

(Poales: Poaceae), the mobility of As and Fe decreased in either shallow or intermittently flooded tailings because of the buffering by different metalliferous carbonates, and accumulation of organic matter, which enable Zn retention in sediments (Jacob & Otte 2004). Reductions in concentrations of N, P, and metals have been shown to occur as water passed through a natural wetland adjacent to an abandoned Pb–Zn mine, operational in the 1880s–1950s (Beining & Otte 1996).

Natural colonisation of metal-contaminated tailings and leachate ponds can be an effective means of metal removal and stabilisation through biogeochemical processes, as well as uptake by plants (Jacob & Otte 1996). The present study was conducted in an Au and Cu mine in central-western New South Wales. Data from

2003 to 2009 pertaining to leachate exiting the toe of the mine’s south waste-rock dump (SWRD) show the mean pH was 7.9 with elevated levels of Cu, Mn, and Zn.

Non-acidic leachate can include elevated levels of dissolved metal concentrations

(Plumlee et al 1999). The leachate ponds at the study site have been naturally colonised by Typha domingensis, establishing large single-species stands. Pre- mining vegetation surveys of this area (1987–1995) list T. domingensis in waterways Bower & Medd 1995). T. domingensis is an emergent perennial known to inhabit heavy-metal contaminated waterways (Ye et al. 1998) and storage ponds at mine sites (Dunbabin & Bowmer 1992). Typha is known for its capability to tolerate contaminated aquatic environments (Boers & Zedler 2008) and also to perform well in aquatic environments contaminated by P (Miao & DeBusk 1999),

2- Hg (Arfstrom et al. 2000), and SO4 (Li et al. 2009). It survives and performs in P,

2- Hg, and SO4 contaminated sites by creating an oxidised rhizosphere, which mobilises O2 to its roots (Wang et al. 2008). Species of Typha are preferentially used in constructed wetlands (CWs) for decontamination of waterways (Maine et al

2009; Vymazal 2011). Against such a background, the purpose of this study was to verify the performance of naturally colonised populations of T. domingensis growing in waste-rock dump leachate ponds at a Au–Cu mine site in central- western NSW, by evaluating concentrations of Cu, Mn, and Zn in the leachate, sediment, and roots and shoots of T. domingensis. These metals occur in greatest concentrations in leachate and are therefore targeted in the present study. The study’s purpose was further reinforced by determining levels of metal concentrations in the flowing leachate after it passed through the naturally occurring populations of T. domingensis in the waste-rock dump leachate ponds.

2.2 Materials and methods

2.2.1 Study sites

The study site is 25 km south-west of Orange in the Central Tablelands of New

South Wales (33°30’ E, 148°59’ N; 750 m above sea level) (Figure 4). Mining operations commenced in August 1998, following the discovery of low-grade Au ore in 1992. Low-grade Au–Cu ore is extracted from the Cadia Hill open pit and high-grade Au and Cu ore from the Ridgeway and Cadia East underground mines, within the boundaries of the current mining lease. The Au–Cu mineralisation is hosted by sheeted quartz veins and sheeted quartz sulphide veins that occur in

Ordovician volcanics and sediments (MESH 2009). Waste rock is placed in the

SWRD, which occurs south of the open pit and spreads over 442 ha and contains approximately 430 mt of excavated material. The present study focuses on the leachate draining from the SWRD. The leachate exiting the SWRD is stored in

NLP and SLP, on the western side of SWRD. The leachate depth within each pond is deliberately kept constant (approx 4 m at the deepest point) by mine engineers, pumping the excess leachate to an on-site mill for further use in ore extraction. The leachate flows from the toe of the waste-rock dump along a natural drainage line

(100 m long, 2 m wide, 1 m deep) to the NLP and SLP. A third sampling site, nominated as a reference site, was established upstream of the mine on

Cadiangullong Creek (CAC), a nearby first and second order upland stream that flows through the mining lease and draining into the Belubula River. The mining company has constructed a dam on CAC upstream from mining activities to provide a water source for ore extraction purposes (King et al. 2003). The CAC varies in width from 0.5 m to 5 m and mean depth is 2 m.

Typha domingensis covers~0.17 ha, at NLP and SLP. At the NLP, leachate passes through the T. domingensis stand before reaching the leachate pond, whereas at

SLP, leachate resides at the toe of the dump, where a large stand of T. domingensis occurs, and exits at the southern end of the stand passing through several <5m2 patches of T. domingensis, before entering and settling in the leachate pond. At the

CAC site, T. domingensis occurs scattered along the creek in stands occupying areas ranging from 1 m2 to 5 m2. At SLP, NLP, and CAC, five sampling points at

10 m intervals were nominated along a 50 m transect, between the waste-rock dump and leachate pond at NLP and SLP (Figure 3). At CAC, samples were also collected along a 50 m transect below the dam wall on CAC. Sampling was done to obtain data on the levels of Cu, Mn, and Zn as the leachate travelled through naturally occurring populations of T. domingensis and away from the toe of the waste-rock dump to the pond, and at CAC to record natural concentrations in leachate, sediment and plant material. Leachate, sediment, and root―shoot samples were collected along the transect at each site. Leachate was collected in 1 L polyethylene bottles. Sediment samples were collected at approximately 20 cm depth using an Undisturbed Wet Sampler (Model UWS35, Dormer Engineering

Products, Murwillumbah, Australia) and the sample cores were collected at a depth of 15 cm in 4.4–cm wide removable plastic tubes. Sample material was collected over two seasons winter 2010 (11–14 July) and early autumn 2011 (13–16 March).

From each sampling point, 1–3 entire plants of T. domingensis were collected from a 2x2 m2 plot, along with surrounding leachate and sediment. Determination of T. domingensis was done by referring to (Briggs & Johnson 1968).

1 Leachate outlet 2 3 4 5 SWRD

Figure 3. Arial image of sampling points (dots) within T. domingensis stand in NLP (northern leachate pond) at the foot of the SWRD (outlined in black) (Scale: 1 cm = 23 m).

2.2.2 Sample preparation and analysis

Leachate samples were sterilized through filtration at 0.45 µm with sterile-syringe driven filter units (Millex®―HP, Carrigtwohill Co, Cork, Ireland) before analysis.

Each core sample of the sediment was divided into material from 0–5 cm and 5–15 cm depths, oven dried at 70°C for 48 h, and pulverized in a hand-held mortar. The pulverized sediment was passed through a 500 μm sieve (Endecotts Ltd, Laboratory

Test Sieve, London, ISO 3310―1:2000). For total metals, sediment subsamples

(0.5 g) were digested with 10 mL HNO3 and 1 mL HCI at 121°C until a colourless liquid was obtained. After cooling, the solution was adjusted to 50 mL with deionised water. For metals extractable in diethylene-triamine-pentaacetic acid

(DTPA), 5.0 g sediment was added to 12 ml DTPA solution (0.005 M DTPA, 0.01

M CaCI2, 0.10 M triethanolamine [TEA], pH 7.3) (Lindsay & Norvell 1978).

CAC N

Open-pit

NLP

SWRD SLP

Tailings dam

Plant samples were separated into roots and shoots and oven dried at 70°C for 48 h.

Dried plant samples were then pulverized in a plant grinder with a 0.5 mm mesh screen (J.P. van Gelder & Co. Pty Limited, Woy Woy, Australia). Subsamples (0.5 g) were digested with 10 mL HNO3 at 60°C for 18 h, after cooling the suspension was adjusted to 50 mL with deionised water. Samples of leachate, sediment and 70

plant roots and shoots were analysed for Cu, Mn, and Zn using an ICP–OES (710

ES, Varian 710 ES, California, USA). Each sample was digested as three replicates.

Accuracy and precision of the digestion procedure and analysis was verified with reference material CRM024–050 (Pasture grasses) and ASP 44-08 (Loamy sand) reference material. Plant root samples were also visually identified for the presence of any reddish-brown coloured coatings, typical of Fe root plaques indicating the likely presence of Fe oxyhydroxides (Taylor et al. 1984).

2.3.3 Data analysis

Leachate-sampling data collected in winter 2010 and early autumn 2011 were analysed using curvilinear regression to determine an exponential response for concentrations of Cu, Mn, and Zn to evaluate any reduction in metal concentrations in the leachate as it passed through the T. domingensis stands. Concentrations of

Cu, Mn, and Zn in sediment with DTPA-extractable and the total, acid digested were subjected to a one–way analysis of variance (ANOVA). Generated means are tabulated. All data were analysed using GenStat 14.2 (2011). Determination of metal translocation from roots to shoots was done with the translocation factor (TF) by expressing the ratio of [Metal]Shoot/[Metal]Root (Stoltz & Greger 2002).

2.4. Results

2.4.1 Leachate and creek water analysis

Concentrations of Cu, Mn, and Zn in leachate declined significantly (p<0.05) from the toe of the waste-rock dump to the leachate-collection ponds in early autumn

2011 (Figure 5 A–C). Concentrations of Cu, Mn, and Zn in the leachate did not decline in locations between the toe of the waste-rock dump and leachate ponds in winter (data not shown).

2.5 Sediment analysis (DTPA extraction, 0–5 cm and 5–15 cm depths)

Mean concentrations of DTPA-extractable Cu and Zn were the highest at 0–5 cm depth at the SLP, with 200 mg/kg for Cu and 30 mg/kg for Zn, whereas the highest

Mn concentration was at CAC with 384 mg/kg from the winter 2010. This trend also occurred in early autumn 2011 sampling at SLP, although the concentrations were less with Cu 103.2 mg/kg and Zn 17.6 mg/kg, and Mn at CAC 131 mg/kg.

Mn concentrations at 5–15 cm depths were the highest at CAC with 312 mg/kg in the winter 2010, whereas concentrations of Cu and Zn were the highest at NLP with 76 mg/kg and 11.4 mg/kg in winter 2010. For early autumn 2011 Cu concentrations were the highest at SLP with 61 mg/kg and Zn at CAC with 7 mg/kg, Mn concentration remained the highest at the CAC with 133 mg/kg.

2.6 Sediment analysis (Total acid extraction, 0–5 cm and 5–15 cm depths)

Mean concentrations of total Cu and Zn in the 0–5 cm depth were the highest at the

SLP with 2042 mg/kg and 480 mg/kg, Mn concentrations in the 0–5 cm depth were the highest at 5529 mg/kg at CAC from winter 2010 sampling. For the early autumn sampling Cu concentrations were the highest at NLP with 857 mg/kg, Mn was highest at SLP at 1297 mg/kg along with Zn at 142 mg/kg. Mean concentrations of total Cu and Zn at 5–15 cm depth were the highest in the NLP with 857 mg/kg 132 mg/kg and the highest total mean concentrations of Mn at the

5–15 cm depths were at CAC as 1456 mg/kg in winter 2010 sampling. For early autumn 2011 sampling, Cu and Zn concentrations were the highest at the NLP with

843 mg/kg and 140 mg/kg. Mn concentrations were highest at CAC with 944 mg/kg.

2.7 Plant analysis (Roots and shoots)

Mean root accumulation of Cu and Zn by T. domingensis was the highest at the

SLP at 322 mg/kg and 179 mg/kg, Mn accumulation was the highest at the CAC with 4276 mg/kg in winter 2010 sampling. Root accumulation of Cu and Zn in winter 2010 was the highest at the SLP with 225 mg/kg and 101.6 mg/kg, accumulation of Mn was the highest at CAC with 1932 mg/kg. Mean-shoot accumulation of Cu by T. domingensis was the highest at the SLP with 83.2 mg/kg in winter 2010, while the highest accumulation of Mn and Zn was in early autumn

2011 sampling with 2324 mg/kg at CAC and 55.3 mg/kg at the SLP. Root samples 74

of T. domingensis collected from CAC displayed a reddish-brown colouring, while this colouring was not apparent on those collected from NLP (Figure 6. A–B).

(A)

(B)

Table 6. Mean values of Cu, Mn, and Zn (mg/kg) in sediment from DTPA-extractable and total (acid digested) at 0-5 cm and 5-15 cm depths and root and shoots of T. domingensis at northern leachate pond (NLP), southern leachate pond (SLP), and Cadiangullong Ck (CAC) during early autumn (13-16 March, 2011) and winter (11-14 July, 2010) sampling, n=6.

Sample material Depth Site Winter sampling (11-14 July, 2010) Early autumn sampling (13-16 March, 2011) (cm) Cu (mg/kg) Mn (mg/kg) Zn (mg/kg) Cu (mg/kg) Mn (mg/kg) Zn (mg/kg) Sediment (DTPA) 0-5 CAC 12a 384a 12.6a 7.1a 131a 7.0a NLP 49a 7b 2.9a 42.1a 7b 2.7a SLP 200b 17b 30a 103.2b 39bc 17.6b Sediment (DTPA) 5-15 CAC 9a 312a 11.4a 7a 133a 6.9a NLP 76a 8b 3.4a 38a 7b 2.3a SLP 71a 6b 10.3a 61a 23b 6.8a Sediment (total) 0-5 CAC 53a 5529a 123a 51a 1072ab 94.9a NLP 1175b 976b 122a 857b 680a 141.5b SLP 2042b 3547a 480b 538c 1297b 142b Sediment (total) 5-15 CAC 42a 1456a 120a 48a 944a 97.2a NLP 1384b 1115ab 132a 843b 696a 140.6b SLP 488c 724b 132a 393c 863a 105.1a T. domingensis Roots CAC 9a 1932a 14.6a 30a 4276a 41a NLP 60a 70b 27.3a 219b 176b 112b SLP 225b 878ab 101.6b 322b 814b 179c T. domingensis Shoots CAC 3.3a 1265a 6.6a 21.5a 2324a 20.1a NLP 36.4b 135b 21.3b 66.6b 551b 45.2b SLP 83.2c 644b 35.7b 47.6c 552b 55.3b Within the different sample materials and sampling times means values of Cu, Mn, and Zn for the different sites with the same letters (superscript) differ significantly (p<0.05).

2.8 Translocation factor

The translocation factor (TF), the ratio of metals stored in shoots vs. roots, indicates internal metal transportation (Keraita et al. 2010); translocation factor

(TF) was calculated as a measure of ions Cu, Mn, and Zn transport from roots to shoots: TF% = (ion content in shoot/ion content in root) X 100. The data presented indicate a strong tendency for translocation of Mn at the NLP site in both winter

2010 and early autumn 2011, whereas Mn translocation was much less at CAC and

SLP. For Cu, higher TF values were obtained in winter 2010 at NLP and SLP, whereas the highest Cu TF value occurred in early autumn 2011 at CAC. Highest

TF values for Zn occurred at NLP in winter 2010.

Table 7. Translocation factors in T. domingensis at Cadiangullong Creek (CAC), northern leachate pond (NLP), and southern leachate pond (SLP).

Site Winter sampling Early autumn sampling (11-14 July, 2010) (13-16 March, 2011) (Metals) (Metals) Cu Mn Zn Cu Mn Zn CAC 0.36 0.65 0.45 0.71 0.54 0.49 NLP 0.60 1.92 0.78 0.30 3.13 0.40 SLP 0.36 0.76 0.35 0.14 0.67 0.30

2.9. Discussion

Concentrations of Cu, Mn, and Zn in leachate were all significantly reduced at the

NLP during the summer. These findings support the capacity of natural processes to contain the spread of contamination and reduce the concentrations of pollutants at contaminated sites (Mulligan and Young 2004). Naturally colonised plant

populations growing in mine tailings has been reported to be an effective means of reducing concentrations of metals in leachate. Effectiveness of Pb and Zn retention by naturally colonised populations of Glyceria fluitans in mine tailings pond in the

Wicklow Mountains National Park (Ireland) demonstrates the capability of natural remediation processes (Jacob & Otte 2004). Additionally, immobilisation of metals from alkaline-mine drainage was achieved in a natural wetland surrounding a mine site in the Farr Creek Drainage area in Cobalt, Canada with up to 20% of the total mass of metals removed by naturally colonised populations of T. latifolia and the remainder through sediment accumulation (Kelly et al. 2006). Treatment of acidic

(pH 3.0) mine leachate by a natural wetland with populations of P. australis in southern Hokkaido (Japan) reduced metal concentrations in leachate along a 1.2 km distance as follows: Cu 1.7 mg/L; Mn 8.5; Zn 4.5 (Sasaki et al. 2003).

Total and DTPA-extractable concentrations of Cu and Zn in sediments were the highest in NLP and SLP, whereas Mn was highest at CAC (Table 7). Additionally, a strong seasonal variation in sediment concentrations of Cu, Mn, and Zn were observed, with greater levels occurring in the winter, compared with summer. A seasonal variation of chemical composition in water and sediments was reported by

(Shomar et al. 2005), although the concentrations were elevated in summer, due to human inputs, while concentrations were less in winter due to precipitation inputs.

(Wright & Mason 1999) reported greater metal concentrations in sediment during summer compared with winter in wetlands in eastern England, and linked their findings with (Thompson et al. 1984), stating that greater concentrations in summer

were due to increased run-off and sewage. Metal concentrations in the sediment in the present study oppositely align with those reported earlier, since greater concentrations occurred in winter 2010 and less in summer 2011. Metal concentrations in leachate at NLP and SLP were greater during the summer but declined significantly at NLP. This finding highlights that metal concentrations in the leachate were low along the transect, and subsequent accumulation of metals from the sediment by T. domingensis in high likelihood contributed to a reduction in metal-sediment concentrations in summer, due to the decrease of metals in leachate along the NLP transect. Seasonal variations of metal levels in plants are known in several species used as remediating agents in metal-contaminated sites

(Weis & Weis 2004). Concentrations of Zn, Pb, Cu, and CD in Spartina maritima

(Poaceae) roots vary seasonally, with low levels in winter and greater levels in spring and summer in Portugal; the greatest increase of metals occurred at the beginning of the growth period (March), which could be due to modification of the chemistry of the rhizo-sphere during active-growth periods of plants, which consequently alters the metal availability at this interface (Caçador et al. 2000).

Metal concentrations in shoots are known to be higher in autumn, due to an increase in aboveground tissue over the active growing period; as this period is followed by a senescent period plants may use this as a detoxification mechanism as metals are shifted from roots to shoots and released during the winter months

(Weis et al. 2003).

The presence of T. domingensis is considered a vital component of the studied wetland system, contributing to immobilisation of metal impacted waters and sediment. The elements Cu, Mn, and Zn were selected for analysis due to the higher concentrations in the mine water at NLP and SLP. Manganese in particular, is often known to be the most toxic metal in mine water discharge (Benner et al.

1995).However, the results of this study revealed significantly greater concentrations of total and bio-available Mn in sediment at the CAC site. These findings occurred despite the significantly lesser concentrations of Mn in the creek water at CAC compared with the leachate at NLP and SLP. This indicates that various plant and sediment processes were responsible for greater accumulation and retention of Mn in a waterway not impacted by mining activities. In line with the findings of significantly greater concentrations of Mn in sediment; accumulation of

Mn by T. domingensis was significantly greater in roots of plants collected from the

CAC site.

Visual assessment of roots of T. domingensis growing at all sites was visually examined for root plaques, indicating the incidence of FeO(OH). Root samples from CAC showed distinct reddish–brown colouring, typical of Fe root plaque

(Batty et al. 2002). Fe-root plaques adsorb and co-precipitate Mn (Ye et al. 1998),

The reported Mn concentrations in roots of T. domingensis at CAC are likely due to co-precipitation with Fe root plaques (Caçador et al. 2000). Typha has well developed air chambers that enable oxygenation of anaerobic substrates (Brix et al.

1992). In addition to the role oxygenation of by roots of T. domingensis, the input

of streamflow further stimulates the rate of manganese oxidation causing dissolved

Fe and Mn to become oxidised and precipitate (Wielinga et al. 1999). Oxidation of the wetland sediments stimulates the rate of manganese oxidation, producing new sorption sites in the form of manganese oxide coatings which enhances removal of trace metals (Fuller and Harvey 2000). It is therefore evident that the sampled conditions of the CAC are more consistent with a greater velocity of streamflow compared with the NLP and SLP sites.

The absence or presence of Fe root plaques has been shown to reduce the movement of metals from roots to shoots; furthermore the mobility of metals in wetland plants is known to be restricted to the root-zone (Wu et al. 2013).

Translocation of metals from roots to shoots appeared to be influenced by the presence of Fe-root plaques restricting shoot translocation of Mn at CAC, while at

NLP shoot translocation was enhanced. Accumulation through above ground tissue via the leaf surface has been reported to greatly assist in shoot accumulation

(Levine et al. 1990). In the present study all plant samples were washed before analysis, however the direct absorption of airborne metals throughout the life history of the sampled plants within an active mine-site could not be totally eliminated (Wu et al. 2013).

Accumulation of Mn by T. domingensis at CAC was the greatest in early autumn

2011, since plant growth increased through summer: a greater quantity of oxygen would have been provided to the root zone (Armstrong & Drew 2002) enhancing

Fe root plaque and Mn co-precipitation. Fe root plaques also play a role in metal translocation and act as a filter for the movement of Fe, Cu, Cd, Ni, and Zn (Liu et al. 2004). Fe root plaques were not evident on roots in NLP in winter 2010 and early autumn 2011, and the greatest translocation factor (TF) factor for Mn occurred at NLP in early autumn sampling with a TF of 3.13 (Table 8). Despite high rates of Mn concentrations in sediment and roots in CAC, the TF was much less compared with NLP (Table 8) indicating absence of Fe root plaques on T. domingensis roots may have contributed to a greater shoot translocation at NLP.

However, Deng et al. (2004) have reported low TF rates when metal were more abundant in the sediment and greater rates when metal concentrations in sediment were less. Although, concentrations of Cu in shoots of Phragmites australis were less in the presence of Fe root plaque (565 mg/kg) than when it was absent

(1400mg/kg) (Batty et al. 2000). Fe root plaques can also have a high affinity for absorption and co-precipitation of Zn (St-cyr & Campbell 1996). Sediment concentrations of total Zn were the highest at the SLP in winter with 480 mg/kg.

The highest accumulation of Zn by T. domingensis also occurred at SLP, in winter with 179 mg/kg for roots and 55.3 mg/kg for shoots. The TF for Zn was consistent at each site and season with TF in the range of 0.30–0.49, with the exception of the

NLP during the winter sampling with a TF of 0.78.

Chapter 3: Hydroponic assessment of four emergent wetland species for phytoremediation of wetlands impacted by saline-mine leachate

3.1 Introduction

The mining industry in Australia in particular is currently experiencing growth due to the rising value of metals such as gold, copper, nickel, and zinc. An increase in metal-ore extraction consequently generates large quantities of waste-rock dumps.

Stockpiled waste-rock dumps at mine sites are responsible for the generation of acidic, saline, neutral, and metalliferous leachate due to the incidence of sulphide- based minerals, such as pyrite (Plumlee et al. 1999). Saline leachate is derived from waste-rock with low-pyrite content and acid-neutralising carbonates (Brady et al.

1998); saline leachate usually contains elevated concentrations of sulphates of calcium and magnesium and a neutral pH and low metal concentrations (INAP

2009). High concentrations of sulphate occurring in leachate is toxic to soil and water organisms (Guilizzoni 1991). Additionally, greater concentrations of sulphate also damage aquatic-plant root systems (Smolders and Roelofs 1995). Sulphur occurs in plant tissues in varied complex forms: absorbed as SO4s, when

2- translocated and stored in plant tissues, they occur as SO3s, SO2s, and even as S , which are then transformed as sulphur containing organic compounds, such as cysteine, Υ-glutamylcysteine, glutathione, for instance (Hawkesford & De Kok,

2006).

Constructed wetlands for the treatment of contaminated wastewater sites are used extensively throughout the world (Pilon-Smits 2005). Constructed wetlands at mine sites are a reliable option for treating mine-waste leachate (Dunbabin and Bowmer

1992; O'Sullivan et al. 2004; Mayes et al. 2009). These wetlands planted with appropriate aquatic plants facilitate a decrease in metal occurrence by sequestering them in plant parts (e.g., root systems), stabilisation of sediment surfaces, maintenance of flow patterns and the addition organic matter, which also act as a valuable carbon source for microorganisms (Batty 2003). The ability of wetland plants to not only accumulate but also tolerate the impact of high concentrations of chemical elements (metallic/non-metallic) is also paramount for the efficient function of constructed wetlands (Zhang et al. 2010).

Hydroponic-screening experiments to evaluate potential-candidate species provide an ideal platform for selecting candidate plants for phytoremediation projects. A beneficial feature of hydroponic experiments is the ability to expose plants to precise concentrations of the elements to be tested against (Moreno-Jiménez et al.

2010). This method is also a cost-effective alternative against time-consuming and expensive-field experiments (Migeon et al. 2012). Hydroponic experiments running for 14‒21 days provide a dependable indication of the trialled plants’ ability to tolerate greater and varying concentrations of elements (Nyquist and Greger 2007;

Comino et al. 2009; Zhang et al. 2010). In particular, the ‘short-term root- elongation test’ running over 14‒21 days is a widely used test to quantify tolerance for several different metals such as Pb, Zn, Ni, Co, and Cd (Köhl and Lösch 2004).

A variety of wetland species are used in constructed wetlands. The most commonly used are emergent macrophytes, such as species of Typha (Typhaceae) and

Phragmites (Poaceae) (Zhang et al. 2010). Although many aquatic species show promise in the treatment of mine leachate, we selectively explored the suitability of four native-wetland species of Australia. Species selection was made considering their endemicity to central-western New South Wales (Sainty and Jacobs 2003) and those that are growing in mine waters in central western NSW.

Each species was assessed for its photosynthetic performance — an indicator of stress tolerance, and the rates of accumulation of elements in plant parts to determine the phytoremediation capability. The aim of the study was therefore to assess the performance of four native emergent wetland plants in a hydroponic- screening experiment for potential use in the phytoremediation of saline-mine leachate. Specifically, the objectives of this study were to measure (a) photosynthetic-gas exchange rate, (b) accumulation rate of Ca, Cu, Mg, Mn, S, and

Zn in shoots and roots, and (c) growth rate and the translocation parameters such as bioconcentration factor (BCF), translocation factor (TF), and tolerance index (TI).

3.2 Materials and methods

3.2.1 Plant material and treatments

The experiment was conducted in a temperature-regulated glasshouse (25–15°C day–night, 50–70% relative humidity), with natural, ambient-light conditions at 85

Charles Sturt University, Orange, NSW, Australia. Six-month old seedlings of B. articulata, C. appressa, S. validus, and P. australis were obtained from a commercial-plant nursery; after washing the soil adhering to roots with distilled- water, seedlings were transplanted into plastic containers at the rate of one of each seedling/container containing a modified Knop’s nutrient solution [10 mL of 10%

Ca(NO3)2, 3 mL of 10% KNO3, 1 mL of 10% KCI, 10 mL of 2.5% KH2PO4, 5 mL

2- of 5% MgSO4 and 1 mL of 0.25% FeCI3 in 1 L of distilled water adjusted to 5.5 pH] (Mleczek et al. 2010). Following 7 d of plant acclimation in the nutrient solution, the seedlings were exposed to treatment solutions consisting of Ca2+,

Cu2+, Mg2+, Mn2+, S2+, and Zn2+. In total, four treatments (referred hereafter as

Conc. 1—4) (Table 10) were prepared. The control-plant populations were supplied only with the nutrient solution. The nutrient solution plus the added elements

(treatment) and the nutrient solution only (control) were replaced weekly, and the solutions were aerated throughout the entire experiment. The experiment was repeated three times with the trial of each species consisting of treatment (x 4) and control (x 2) totalling six (6) of each species per run (72 plants in total).

3.2.2 Determination of porewater reference values for hydroponic solution

The concentrations of elements used in the treatment solution were determined by referring to concentrations of the trialled elements found within porewater collected from a pond receiving saline-mine leachate from an active Au and Cu mine (Adams et al. 2013). Porewater contains the aqueous bioavailable phase of the trialled elements to which plant roots were exposed (Nolan et al. 2003) and were 86

considered as suitable reference values for a hydroponic experiment to determine tolerance to wetland conditions impacted by saline-mine leachate. Sediment was randomly collected to a depth of 20 cm and immediately returned to the laboratory for extraction of porewater. Porewater was extracted by spinning the samples in a bench centrifuge (MSE Mistral 2000, Leicester, United Kingdom) at 4000 G for 10 min. Centrifuged samples were filtered at 0.45 µm using sterile-syringe driven filter units (Millex®–HP, Carrigtwohill Co, Cork, Ireland). Concentrations of Ca, Cu,

Mg, Mn, S, and Zn were determined in an ICP–OES system (Varian 710 ES,

California, USA). Prior to analysis, the porewater samples were acidified to a pH of

2±0.5 for elemental analysis to ensuring that the elements remained in suspension

(Batty et al. 2006).

Table 8. Porewater concentrations (mg/L) of Ca, Cu, Mg, Mn, S, and Zn from leachate storage pond receiving saline-mine leachate from a gold and copper mine in Central-tablelands, NSW, Australia (Adams et al. 2013). Mean data and SD values presented, n=6.

Element Leachate pond Ca 502.8 ± 32.69 Cu 0.169 ± 0.01 Mg 423.8 ± 29.35 Mn 0.617 ± 0.26 S 2872 ± 67.84 Zn 0.125 ± 0.01

Table 9. Concentrations of elements dissolved in hydroponic solution.

Element concentrations (mg/L) Ca Cu Mn Mg S Zn Conc. 1 500 4 15 500 1977 4 Conc. 2 250 3 7.5 250 989 3 Conc. 3 125 2 3.25 125 494 2 Conc. 4 62.5 1 1.6 62.5 247 1

3.2.3 Growth measurements

After 14 days of treatment, the dry mass (DM), root length, and ion accumulation were determined. Dry masses of whole plants were obtained by desiccating the plant material at 70°C until they stabilised at a constant value. The tolerance index

(TI) was calculated to measure the plant capacity to increase biomass under treatment conditions: TI % = (root length in the treatment)/(root length in the control) X 100. Bioconcentration factor (BCF) was used as a measure of the uptake of Ca, Cu, Mn, Mg, S, and Zn from the treatment solution: BCF = element concentrations in shoots (mg/kg-1) / element concentration in solution (mg/L-1).

Translocation factor (TF) was calculated as a measure of ions Ca, Cu, Mn, Mg, S, and Zn transport from roots to shoots: TF% = (ion content in shoot/ion content in root) X 100.

3.2.4 Gas-exchange measurements

After 14 days of treatment in the hydroponic medium, leaves of each species and the same physiological age were selected for gas-exchange measurements. Net- photosynthetic rate (Pn), stomatal conductance (gs), internal CO2 concentration

(Ci), and transpiration (T) were assessed on intact leaves using a portable photosynthesis system (LI–COR 6400XT, Lincoln, Nebraska), equipped with a clamp-on leaf cuvette that exposed 2 cm2 leaf area. Light and humidity were 500

µmol m–2s–1, 50 % ± 2 %, respectively; the experiment was conducted in April

–2 –1 (2013). CO2 supply was maintained at a constant level of 380 µmol m s using a 88

CO2 injector (LI-6400-01, LI-COR, Lincoln, Nebraska) with a high-pressure liquefied CO2 cartridge source.

3.2.5 Chlorophyll-fluorescence measurements

Leaf-fluorescence measurements were determined using a portable-photosynthesis system (6400XT, LI–COR, Lincoln, Nebraska) equipped with a 6400-40 Leaf-

Chamber Fluorometer (LCF). Efficiency of photosystem II (PhiPS2) (Genty et al.

1989) was determined, followed by non-photochemical quenching (NPQ), leaves were dark adapted with dark adapting clips (LI-COR 9964–091, Lincoln Nebraska) for 2 h before determination quantum efficiency of Fv/Fm.

3.2.6 Element accumulation

Plants were harvested following gas-exchange measurements, rinsed in distilled water, and sorted into shoots and roots. All samples were dried to a constant mass at 70°C for 48 h. Dry-plant samples were pulverized in a plant grinder with 0.5 mm mesh screen (J.P. van Gelder & Co. Pty Limited, Woy Woy, Australia).

Subsamples (0.25 g) were digested with 5 mL HNO3 at 70°C for 18 h, after cooling the suspension was adjusted to 25 mL with deionised water. Concentrations of Ca,

Cu, Mn, Mg, S, and Zn in plant roots and shoots were analysed by ICP–OES (710

ES, Varian, California, USA). Accuracy and precision of the digestion procedure and analysis were verified using reference material (CRM024–050, RTC, Sigma

Aldrich, St Louis, U.S.). 89

3.2.7 Statistical analysis

Concentrations of Ca, Cu, Mg, Mn, S, and Zn in porewater collected from waste- rock dump leachate ponds were determined by generating a mean and standard deviation from the collected samples. Photosynthetic assessments and metal accumulation were subjected to a two-way analysis of variance (ANOVA).

Bonferroni test was used to determine significant differences between individual means. All data were analysed using GenStat 15th edn.

3.3 Results

3.3.1 Growth measurements

A significant (P<0.05) treatment effect on root length (Figure 7) occurred between treatment conc.1 and 2 of C. appressa. Treatment-root length of C. appressa was greater than the control at treatment Conc. 1. A significant (P<0.05) treatment effect on dry mass occurred between each species and treatment (Figure 7). For S. validus treatments 2 and 3 were greater than the control, C. appressa in treatment 2 was greater than the control, whereas B. articulata was greater than the control at treatment 1. The tolerance index % (TI) (Table 11) did not differ significantly within treatments for each species. The bioconcentration factor (BCF) (Table 12 a– d) differed significantly (P<0.05) between treatments and between each species.

The BCF values for B. articulata, C. appressa, S. validus decreased as follows: 90

Mn>Ca>Mg>Zn>Cu>S, the BCF for P. australis was Ca>Mg>Mn>Zn>Cu>S.

Translocation factor (TF) differed significantly (P<0.05) within each species and

between treatments, B. articulata stored greater quantities of Ca, Cu, Mg, and Zn in

roots in greater treatment solution and translocated more Mn and S to shoots in

greater treatment solution. Carex appressa translocated greater quantities of Mn,

Mg, and S to shoots, whereas translocation of Ca, Cu, and Zn was consistent in all

treatments. Phragmites australis TF was consistent across all treatments; TF for S.

validus behaved similar to that of P. australis with no trends of either root storage

or shoot translocation.

(a) (b)

grams) grams)

Mass ( Mass Rootlength (mm)

Treatment 1 2 3 4 B. articulata 93.4a ± 53.4 91.2a ± 5.1 102.2a ± 8.6 105.3a ± 15.1

C. appressa 153.9a ± 83.3 74.9a ± 18.1 94.2a ± 15.9 136.6a ± 75.9

P. australis 106a ± 15.8 123.2a ± 44.1 99.8a ± 23.2 105.1a ± 35.5

S. validus 92.5a ± 27.5 90.3a ± 21.3 107.2a ± 15.6 73.1a ± 3.5

Table 11 A (BCF) Species Treatment Ca Cu Mn Mg S Zn B. articulata 1 11.61ab ± 0.01 8.85a ± 1.8 21.3abc ± 0.51 19.81abc ± 0.57 2.37abc ± 0.3 23.17ab ± 2.58 2 18.87abcd ± 1.62 12.4abc ± 2.92 28.53abcd ± 3.01 22.32abcd ± 1.79 2.15ab ± 0.82 21.65ab ± 1.48 3 37.38acde ± 0.42 15.82abcd ± 5.62 40abcdef ± 0 38.04abcdef ± 0.65 3.03abcde ± 0.75 23.47abc ± 1.02 4 60.95fgh ± 2.16 25.75abcd ± 3.46 62.15fg ± 3.13 50.14fg ± 0.15 3.3abcde ± 1.6 38.7abcd ± 0.28

C. appressa 1 19.1abcd ± 2.9 11.78abc ± 1.43 29.83abcde ± 4.28 34.22abcdef ± 5.17 4.05abcdef ± 0.33 32.25abcd ± 4.94 2 42.75cef ± 7.38 22.61abcd ± 1.9 73.2g ± 10.55 46.35ef ± 7.61 5.61abcdef ± 0.54 46.33d ± 7.54 3 67.87gh ± 5.02 43.1ab ± 16.12 54.46bdefg ± 5.65 77.18h ± 5.1 6.08abcdef ± 1.67 44.55acd ± 6.85 4 130i ± 9.85 30.9abcd ± 10.04 116.25h ± 7.07 107.45i ± 8.07 9.42acdefg ± 2.75 76.55e ± 2.75

P. australis 1 10.76ab ± n/a* 8.9ab ± n/a* 15.13a ± n/a* 13.57a ± n/a* 1.89a ± n/a* 25.75a ± n/a* 2 16.36abc ± n/a* 10.1ab ± n/a* 20.53ab ± n/a* 17.43ab ± n/a* 2.98abcd ± n/a* 18.83abcd ± n/a* 3 54.18efg ± n/a* 17.9abcd ± n/a* 36.61abcdef ± n/a* 47.6efg ± n/a* 6.37abcdef ± n/a* 28.25abcd ± n/a* 4 89.26h ± n/a* 24.7abcd ± n/a* 63.125efg ± n/a* 72.97fg ± n/a* 12.29dfg ± n/a* 73.1e ± n/a*

S. validus 1 18.7abcd ± 1.23 15.98abcd ± 0.47 33.46abcde ± 1.22 20.58abc ± 1.12 3.37abcde ± 0.16 29abcd ± 1.06 2 32.76abcde ± 0.7 19.85abcd ± 0.07 61.46fg ± 3.95 28.55abcde ± 1.34 4.92abcdef ± 0.81 30.05abcd ±0.07 3 65.84fgh ± 0.84 21.85abcd ± 1.34 105.53h ± 2.61 41.32bdef ± 0.5 7.06abcdef ± 1.44 30.5abcd ± 0.42 4 143.65i ± 9.13 36.25abcd ± 7.28 159.06i ± 11.93 91.09hi ± 5.41 16.17g ± 3.04 79.15e ± 9.26

Table 11 B (TF) Species Treatment Ca Cu Mn Mg S Zn B. articulata 1 78.4ab ± 7.49 9.84ab ± 2.62 45.3ab ± 5.49 121.8abc ± 11.25 129.3abc ± 20.2 26.9abc ± 4.85 2 53.4a ± 4.94 8.5a ± 1.94 22.3a ± 1.9 85.1ab ± 6.89 84.1ab ± 14.14 16a ± 1.21 3 57.5a ± 3.53 6.37a ± 2.62 30.3a ± 2.04 84.9ab ± 3.15 67.6ab ± 0.35 17.78ab ± 1.48 4 194.7def ± 21.9 310.3f ± 43.9 23.7a ± 1.05 486.6cd ± 43.17 10.5a ± 1.94 364.2e ± 50.4

C. appressa 1 132.1abcdef ± 7.83 10.2ab ± 1.03 94.3bc ± 10.34 244.8cdefg ± 31.28 165.6abc ± 18.9 49.3c ± 5.44 2 155.7cdefgh ± 16.78 21.3abc ± 1.75 106.1bcd ± 10.69 231.1cdefg ± 26.66 161.3abc ± 16.18 50.6c ± 5.52 3 114.4abcd ± 9.94 20.4abc ± 7.4 57.61ab ± 5.19 205.8bcdefg ± 18.5 107.2abc ± 17.77 48.3c ± 7.62 4 121.2abcdef ± 8.79 16ab ± 3.49 73.3abc ± 8.72 170.9abcdef ± 13.62 125.3abc ± 9.89 31.2abc ± 6.79

P. australis 1 115.3abcde ± n/a* 11.3ab ± n/a* 74.4abc ± n/a* 144.6abcd ± n/a* 63.2ab ± n/a* 39abc ± n/a* 2 120abcde ± n/a* 16.5abc ± n/a* 38.4ab ± n/a* 141.1abcd ± n/a* 89.7abc ± n/a* 29.2abc ± n/a* 3 182.9bcdefgh ± n/a* 30bc ± n/a* 88.8abc ± n/a* 203.3abcde ± n/a* 89abc ± n/a* 51.3abc ± n/a* 4 107.4abc ± n/a* 19.1abc ± n/a* 46.1ab ± n/a* 169.3 ± n/a* 85.2abc ± n/a* 40.3abc ± n/a*

S. validus 1 210.5cefgh ± 36.59 20.4abc ± 3.16 184.1ef ± 27.51 259.8defg ± 45.53 222.7cd ± 63.9 49.9c ± 8.41 2 240.7h ± 10.06 33.4c ± 0.11 197.4f ± 13.31 247.7defg ± 12.98 215.7cd ± 14.99 49.1c ± 0.07 3 148.4bcdefg ± 7.85 12.7ab ± 1.28 168.6def ± 8.27 194.8bcdefg ± 7.54 178.9bc ± 26.56 31.6abc ± 1.13 4 223.3egh ± 44.41 21.2abc ± 6.53 125.2cde ± 25.6 296.8eg ± 55.6 311.5d ± 29.2 43.6bc ± 9.11

3.3.2 Gas-exchange measurements

The measured net-photosynthetic rate (Pn), stomatal conductance (gs), internal CO2 concentration (Ci), and transpiration (T) (Figure 8) of each species were not significantly (P<0.05) reduced compared with the control. However, a significant- treatment effect (P<0.05) between species occurred. Pn was in the following order:

S. validus>P. australis>C. appressa>B. articulata. The treatment effect on Pn was most evident at Conc.1–2. As concentrations of the elements decreased in the solution, Pn performance increased reaching the greatest rate Conc. 3 for all species. The trends of gs and T aligned closely with Pn. The lowest Ci occurred in

P. australis at Conc. 3, and the highest with B. articulata at treatment Conc. 3.

(a) (b)

mol H

n µmol CO

µmol µmol mol

ol H

T m T

3.3.3 Chlorophyll fluorescence: Fv/Fm, NPQ, and PhiPS2

The quantum efficiency of Fv/Fm, the efficiency of photosystem II (PhiPS2), non- photochemical quenching (NPQ) (Figure 9) were uninhibited (P<0.05) although a significant treatment (p<0.05) effect between species was evident. All recorded values for Fv/Fm fell between 0.76 and 0.82. The treatment values for C. appressa in particular were greater than the control value, a 2.6 % difference between the control and treatment 3 values. Values of PhiPS2 for all tested species were between 0.24 and 0.47. As per the results for Fv/Fm, the treatment values of C. appressa for PhiPS2 were greater than the control; values for S. validus and P. australis were also greater than the control at treatment 1; whereas only the lowest treatment 4 was greater than the control for B. articulata. The treatment effect on

NPQ produced greater values for B. articulata and C. appressa compared with the control, with a 16% and 80% increase compared with the control respectively; while P. australis and S. validus values at Conc. 1 were below the control, at 15% and 44%, respectively.

PhiPS2

NPQ

Figure 9. The effect of treatment in the hydroponic medium on Fv/Fm (A), PhiPS2 (B), and Non-photo chemical quenching (NPQ) (C). Bar indicates l.s.d. at 5 % level, n=6.

3.3.4 Metal accumulation

Mean accumulation of Ca in roots and shoots of each species did not differ significantly among the control and four-treatment solutions. Mean values of Cu accumulation in roots differed significantly (p<0.05) between the control (0) and treatment (Conc. 1–4), although mean-shoot accumulation did not differ significantly between the control and treatment. Mean root and shoot accumulation of Mg differed significantly (p<0.05) between control and treatment 1, mean root and shoot also differed significantly between control and treatment 2 for B. articulata (p<0.05). The highest mean accumulation of Mg was by shoots of C. appressa at 17112 mg/kg. Mean root and shoot accumulation of Mn differed significantly (p<0.05) between control and treatment 1 for all species and treatment

2 for C. appressa roots and shoots and roots only of the remaining species. Mean accumulation of S differed significantly for roots and shoots between control and treatment 1 for C. appressa only and shoots of B. articulata and S. validus, and shoots of C. appressa at treatment 2. Mean root and shoot accumulation of Zn differed significantly (p<0.05) between control and treatment 1 and 2 for C.

appressa, mean root accumulation differed significantly (p<0.05) for all species in all treatment solution with the exception of treatment 3 for P. australis. The highest mean accumulation of Ca, Mg, and S was by C. appressa, while B. articulata accumulated the highest treatment of Cu, Mn and Zn.

Ca Treatment 0 1 2 3 4 B. articulata Shoot 4235 ± 74.2 5807 ± 7.07 4719 ± 406.5 4673 ± 53.7 3809 ± 135.05 Root 7692 ± 911.4 7434 ± 701.4 8827 ± 56.5 8136 ± 593.2 8360 ± 1236.02 C. appressa Shoot 9151 ± 117.3 9552 ± 1460.1 10689 ± 1846.9 8484 ± 627.9 8125 ± 615.8 Root 8026 ± 126.5 7209 ± 677.4 6838 ± 449.01 7415 ± 95.4 6699 ± 21.9 P. australis Shoot 5683 ± 45.2 5469 ± n/a 4179 ± n/a 6860 ± n/a 5666 ± n/a Root 4917 ± 572.7 4959 ± 414.3 3345 ± 83.4 3593 ± 155.5a 4813 ± 536.6 S. validus Shoot 8565 ± 211.4 9350 ± 619.4 8191 ± 177.4 8230 ± 106 8978 ± 570.6 Root 3978 ± 224.8 4483 ± 485 3403 ± 68.5 5548 ± 222 4075 ± 555 Cu Treatment 0 1 2 3 4 B. articulata Shoot 12a ± 4.9 35.4 ± 7.2 37.2 ± 8.7 31.7 ± 11.2 25.8 ± 3.4 Root 68.1abcd ± 14.6 362.5hi ± 23.3 434.5ijk ± 3.5 502.5k ± 30.4 284.5g ± 47.3 C. appressa Shoot 25ab ± 18.1 47.2 ± 5.7 67.9 ± 5.7 86.2bcd ± 32.2 30.9 ± 10 Root 39.3ab ± 15.8 461.5jk ± 9.1 318.5gh ± 0.7 420ij ± 5.6 190f ± 21.2 P. australis Shoot 14.5 ± 3.3 32.6 ± n/a 27.3 ± n/a 32.8 ± n/a 21.7 ± n/a Root 17.7ab ± n/a 2.05 331.5gh ± n/a 24.7 179ef ± n/a 5.6 116cde ± n/a 4.2 118def ± 15.5 S. validus Shoot 8.7 ± 2.82 64 ± 1.90 59.6 ± 0.21 43.7 ± 2.68 36.3 ± 7.28 Root 22.9ab ± 4.52 316gh ± 39.5 178ef ± 0 343.5gh ± 13.4 173.5ef ± 19.09

Mg Treatment 0 1 2 3 4 B. articulata Shoot 2151abc ± 67.8 9908no ± 286.3 5582efghijkl ± 448.3 4755cdefghijk ± 81.3 3134abcdef ± 9.8 Root 3314abcdefg ± 323.1 8157lmn ± 518.3 6559hijkl ± 4.24 5603efghijkl ± 304.05 4649bcdefghijk ± 680.9 C. appressa Shoot 4315abcdefghij ± 53.7 17112p ± 2586.5 11588o ± 1903.5 9647mno ± 638.5 6716ijkl ± 504.8 Root 3120abcdef ± 121.6 6978jklm ± 164.7 4998 ± 246.7 4693 ± 111.7 3928 ± 17.6 P. australis Shoot 3060abcdef ± 38.8 6871jkl ± n/a 4278 ± n/a 5854 ± n/a 4493 ± n/a Root 1912ab ± 1.41 4880cdefghij ± 263.7 3005 ± 115.2 2831 ± 134.3 2625 ± 97.5 S. validus Shoot 3571abcdefg ± 227.6 10294no ± 561.4 7137klm ± 337.2 5165 ± 63.6 5693 ± 338.7 Root 1609 ± 55.8 4004 ± 485.7 2881 ± 14.8 2651 ± 70 1941 ± 249.6 Mn Treatment 0 1 2 3 4 B. articulata Shoot 125.5abcd ± 2.1 319.5hijklm ± 7.7 214.0bcdefgh ± 19.7 130.0abcde ± 0 99.4ab ± 5.02 Root 135.0abcde ± 16.9 705.0r ± 67.8 956s ± 22.6 429.5lmnopq ± 28.9 422lmnopq± 60.8 C. appressa Shoot 141.5abcde ± 0.7 447.5mnopq ± 64.3 549q ± 79.1 177abcdef ± 18.3 186abcdefg ± 11.3 Root 142.5abcde ± 3.5 473.5opq ± 16.2 516pq ± 22.6 307ghijkl ± 4.2 254.5efghij ± 14.8 P. australis Shoot 122 ± 7.07 241.5 ± n/a 143.1 ± n/a 115.1 ± n/a 82.6 ± n/a Root 59.2a ± 2.05 319.5hijklm ± 20.5 390klmnop ± 15.5 130abcde ± 5.65 200.5bcdefgh ± 26.1 S. validus Shoot 374.5 ± 37.4 502opq ± 18.3 461 ± 29.6 343ijklmn ± 8.4 254.5efghij ± 19.09 Root 92.5ab ± 2.8 275fghijk ± 31.1 233.5defghi ± 0.7 203.5 ± 4.9 206 ± 26.8 S Treatment 0 1 2 3 4 B. articulata Shoot 740a ± 656.9 4691efghi ± 609.5 2130abcde ± 813.8 1498abcd ± 374.7 816ab ± 395.2 Root 1011 ± 328.8 3708 ± 1050.7 2486 ± 549.4 2213 ± 542.3 1791 ± 84.8 C. appressa Shoot 1774abcd ± 481.5 8020j ± 654.7 5554hij ± 542.3 3006abcdefgh ± 826.6 2328abcdef ± 680.9 Root 1622abcd ± 354.2 4851fghi ± 158.3 3443 ± 9.1 2776 ± 310.4 1885abcd ± 692.2 P. australis Shoot 3357 ± 567.8 3054 ± n/a 2741 ± n/a 2546 ± n/a 2428 ± n/a Root 3177 ± 196.5 5232 ± 993.4 3079 ± 303.3 2932 ± 856.3 2953 ± 864 99

S. validus Shoot 2618abcdefg ± 599.6 6681ij ± 321.02 4872 ± 810.3 3491bcdefgh ± 713.4 3994 ± 753.06 Root 1267 ± 94.7 3150 ± 1047.9 2251 ± 219.2 1943 ± 110.3 1299 ± 363.4

Zn Treatment 0 1 2 3 4 B. articulata Shoot 46.6 ± 14.2 92.7 ± 10.3 65 ± 4.4 47 ± 2.05 38.7 ± 0.2 Root 85abc ± 9.8 346lm ± 24.04 405m ± 2.8 264.5hijkl ± 10.6 325.5klm ± 50.2 C. appressa Shoot 38.4a ± 1.4 129bcdef ± 19.7 139cdef ± 22.6 89.1 ± 13.7 76.5 ± 2.7 Root 53.3ab ± 4.2 261hijk ± 11.3 273.5ijkl ± 14.8 184.5efgh ± 0.7 252hijk ± 63.6 P. australis Shoot 67.5 ± 7.9 116 ± n/a 50.6 ± n/a 51.6 ± n/a 60.7 ± n/a Root 69.5abc ± 0.3 277jkl ± 18.3 187fgh ± 8.4 105abcde ± 7.07 168.5defg ± 17.6 S. validus Shoot 44.8 ± 2.2 116 ± 4.2 90.2 ± 0.2 61 ± 0.8 79.2 ± 9.2 Root 92abcd ± 2.6 235ghij ± 31.1 183.5efgh ± 0.7 193fghi ± 4.2 183efgh ± 16.9

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3.4 Discussion

Screening for new accumulator plants is needed for phytoremediation in different conditions, new studies are required to find new accumulator plants for use in different climates (Chehregani et al. 2009). The aim of this study therefore was to conduct a short-term screening experiment to select suitable species for use in phytoremediation of saline mine-leachate characterised by excess concentrations of

Ca, Cu, Mg, Mn, S, and Zn.

Wetland plants have a tendency to store greater concentrations of metals in belowground tissues (Batty and Younger 2004), and this trend was particularly evident for P. australis. Restriction of upward movement of metals into shoots

(viz., avoidance mechanism) and translocation of excessive metals to senescing leaves are the key tolerance mechanisms in wetland plants (Weis and Weis 2004).

Although translocation of metals to senescing leaves has been suggested as a viable mechanism for eliminating the burden of metals (Bragato et al. 2006). When grown in a sediment medium, the translocation of elements from roots to shoots is determined by anatomical, biochemical, and physiological features (Salt et al.

1995) and the soil-particle size, organic- matter content, and nutrients (Yang and

Ye 2009). Therefore, a difference of growing media (soil and hydroponic solution) could generate differing results of element accumulation and translocation in the same species. For example, Stoltz and Greger (2002) found translocation factors differed in Eriophorum angustifolium with a greater tendency for the accumulation of Cu, Pb, and As in the shoots in field experiments, but did not show any shoot- 101

accumulation of Cu, Pb, and As in the same species in hydroponic screening.

Hydroponic screening proved useful in determining the response behaviour of Salix genus (Salicaceae) to heavy metals under various long-term field conditions

(Watson et al. 1999). For P. australis, Stoltz and Greger (2002) reported corresponding accumulation traits of Cu between field and hydroponic accumulation, in spite of a greater level of accumulation in roots with hydroponic experiment with 485 mg/kg in roots and 18.7 mg/kg in shoots and in field grown plants 6.4 mg/kg in shoots and 80.1 mg/kg in roots. The cited assessments of P. australis and the present experimental work indicates that P. australis bears a greater affinity to store Cu, Mn, and Zn in belowground tissues than in aboveground tissues (Bragato et al. 2006). However, C. appressa showed greater concentrations in shoots compared with roots. Metal concentrations in plants vary according to the plant part and element (Stoltz and Greger 2002). Phragmites australis has been identified as an excluder of metals from shoots, while the present study has shown that C. appressa in particular translocated significantly greater concentrations of Mg, Mn, S, and Zn when compared against the other tested species. Root uptake with subsequent translocation to above-ground tissues is reported as the main route for aquatic plants (Jackson 1998). The different concentrations of elements in shoots of P. australis and C. appressa may therefore reflect the different pathways for uptake (Baldantoni et al. 2004). The efficiency of excluding metals from shoots can be reduced when photosynthetic activity is reduced or during senescence (Bragato et al. 2006). Photosynthetic activity of C. appressa was reduced but not significantly, and no plants entered a senescent

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phase. It is therefore likely that C. appressa possesses the translocation traits of metal indicators, metal indicators accumulate metals in their above-ground tissues, and the levels in the tissue can reflect metal levels in the soil (Memon and Schröder

2009). In the present study, decreasing BCF values occurred as element accumulation increased in each species and the BCF values at the highest treatment were all >1, a BCF value <1 is a characteristic of metal- excluder species (Baker and Brooks 1989). Bioconcentration values indicate the species’ propensity to accumulate large quantities of metals.

In our present trials, root length and biomass were not significantly impacted.

Although, biomass of Hordeum vulgare and Secale cereale (Poaceae) and Lupinus angustifolius (Fabaceae) declined when growing in alkaline saline-mine tailings in a gold mine in southern New Zealand (Mains et al. 2006). The ability of B. articulata and S. validus to tolerate heavy metals was confirmed by Zhang et al.

(2010), where up to 20 mg L-1 of Cd was supplied in a hydroponic medium.

Photosynthetic performance in the tested plants was not significantly inhibited, indicating that the photosynthetic system in each species was able to tolerate the concentrations of elements in plant tissues. The above-optimal concentrations of each of the tested elements can damage the photosynthetic mechanism. When grown in hydroponic solution containing 300 mg/L of Ca, the net-photosynthetic rate and stomatal conductance of Camellia sinensis (Theaceae) was inhibited

(Wang et al. 2010). All species in the present trial tolerated Ca in excess of 3809 mg/kg in leaves. Excess accumulation of Cu in leaves is attributed to an overall

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inhibition of photosystem II due to chloroplast-membrane leakage, which is a key stress response (Ouzounidou 1994). The wetland plant Spartina densiflora

(Poaceae) tolerated 320 mg/kg in shoots before quantum efficiency of PSII, net photosynthetic rate, and stomatal conductance were inhibited (Mateos-Naranjo et al. 2008). Shoot accumulation of Cu did not exceed 86.2 mg/kg and this occurred in

C. appressa. Mean accumulation of Cu by B. articulata grown in Hoagland medium was 63 mg/kg in roots and 5.8 mg/kg in shoots (Zhang et al. 2010).

Phragmites australis is an established root accumulator of Cu; plants grown at treatment 1 containing 4 mg/kg of Cu accumulated a greater percentage in roots –

331. 5 mg/kg compared with shoots – 32.6 mg/kg. A similar trend was evident when samples of P. australis collected from lagoons receiving mine drainage rich in Cu were trialled in hydroponic medium with 0.5 µg ml-1 Cu; root accumulation was 4186 µg g-1 and shoots 386 µg g-1, field (Ye et al. (2003). The highest mean accumulation of Cu in B. articulata was 502.5 mg/kg in roots at treatment 3 and

37.2 mg/kg in shoots at treatment 2. Excess accumulation of Mn has been attributed to injury in plants, although the tolerance to excess of this metal is highly dependent on the plant species and even genotypes (Foy et al. 1988). Lolium perenne accumulated 2357 mg/kg Mn and suffered chlorotic leaves (Rosas et al.

2007); whereas CO2 assimilation and stomatal conductance of Citrus grandis decreased significantly with 583.6 µg/g (Li et al. 2010). Carex appressa accumulated the greatest quantity of Mn at treatment 2 at 549 mg/kg. Accumulation of S in leaves ranged from 816 mg/kg for B. articulata at treatment 4 to 8020 mg/kg in C. appressa at treatment 1 containing 4 mg/L of S; S. Lycoperiscon

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esculentum was grown in hydroponic medium containing 20.8 mM sulphate significantly reduced photosynthetic capacity and Rubisco activity (Xu et al. 1996).

The toxic effects of Zn were observed in Populus deltoides grown in hydroponic solution containing 5 mM zinc chloride; net photosynthesis was reduced from

15.74 to 3.78 µmol m-2 s-1 (Fernàndez et al. 2012).

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Chapter 4: Carbon assimilation and porewater interactions of macrophytes for phytoremediation of saline mine leachate

4.1 Introduction

Saline-mine leachate is generated in mine waste-rock dumps due to the acid- generating sulphide mineral pyrite; the presence of acid-neutralising carbonates assist in neutralising the pH (Plumlee et al. 1999). Saline mine leachate contains elevated concentrations of Mn, Se, Na, Ca, Mg, NO3, and S (Naftz and Rice 1989).

Abiotic stress induced by heavy metals can negatively impact photosynthetic performance; chromium (Cr) was responsible for a reduction of CO2 fixation in

Helianthus annuus (Asteraceae) (Davies et al. 2002), and Zn inhibited photosynthetic performance in Datura metel, D. innoxia, D. sanguine, and D. tatula

(Solanaceae) (Vaillant 2005). Net-photosynthetic performance and physiological parameters are proven indicators of stress and tolerance to excess concentrations of heavy metals and nutrients (Adhikari et al. 2010). Chlorophyll-fluorescence measurements provide an insight into the ability of plants to tolerate environmental stresses and the extent of damage to the photosynthetic apparatus (Maxwell and

Johnson 2000). Detection and quantitation of stress-induced changes in photosystem II (PSII) can be determined by analysis of chlorophyll fluorescence

(Mehta et al. 2010); light- and dark-adapted measurements of leaves can accurately determine whether photodamage has occurred (Naumann et al. 2008).

Photosynthetic assimilation of CO2 and its partitioning along with N uptake can also be determined by measurement of stable-isotope ratios. Isotopic discrimination

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by plants has been widely utilised to investigate response of plants to the environment (Cabrera-Bosquet at al. 2009; Caldelas et al. 2011). For example, assessment of δ13C composition in leaf tissue provides and an integrated measurement of the ratio of intercellular CO2 to ambient CO2 (C1 - intercellular

CO2/Ca – ambient CO2 concentration), which provides an indication of stomatal limitation of photosynthesis (O’ Leary 1988). The enzyme Ribulose-1,5- biphosphate carboxylase/oxygenase (Rubisco), is responsible for CO2 fixation and is also known to be sensitive to a range of abiotic stress factors including heavy metals (Galmés 2013). In particular Rubisco activity is negatively affected after exposure to Cd concentration above 50 µM (Ying et al. 2010).

The biogeochemical reactions in the sediment environment including mineral precipitation and dissolution reactions govern the mobility of metals in wetland sediments (Moreau et al. 2013). Aquatic plant species such as Phragmites australis and Typha domingensis are able to scavenge and accumulate metals before sediment sequestration (Peltier 2003), through sulfide oxidation and dissolution by radial-oxygen loss (ROL) from roots (Armstrong and Armstrong 1990). In many wetland plants the aerenchyma is well developed even in drained sediments and is further enhanced in waterlogged conditions (Justin and Armstrong 1987).

Aerenchyma provides a low-resistance pathway for the movement of O2 between shoots and roots (Li et al. 2011).

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Table 13. Concentrations of elements in acid- mine drainage and saline mine- drainage (Plumlee et al. 1999).

Element Acid mine-drainage (pH -1–6.2) Saline mine-drainage (pH 5.9 –8.9)

Zn, Cu, Cd, Pb, Co, Ni 10 – 100 000 000 1 – 1 000 000

The release of ROL from roots of wetland plants is an adaptation to overcome anoxia induced by waterlogging; diffusion of oxygen from roots into adjacent soil enables plants to convert soil toxins via chemical, enzymatic and microbial oxidation (Armstrong 1975). The release of oxygen via the root surface and subsequent formation of an oxidative layer on plant roots prevent the plant from absorbing phytotoxic reduced compounds such as Fe2+ and Mn2+ and sulphides, which accumulate in the anoxic wetland sediments (Christensen et al. 1994). The release of oxygen and exudates from roots can result in the formation of iron (oxy- hydroxide) precipitates (St Cyr et al. 1993). Iron coatings on roots consist of lepidocrocite (ᵧFeOOH) (Bacha and Hossner 1977), and goethite (α-FeOOH) (Chen et al. 1980a) and ferric phosphate (Snowden and Wheeler 1995). Root plaques can adsorb and co-precipitate potentially phytotoxic elements, resulting in a depletion of metals from interstitial waters (porewater) (Batty et al. 2006).

Determination of the influence of wetland plants on the mobility of heavy metals in sediments has included in-situ (Wright and Otte 1999) and ex-situ (Batty et al.

2006) assessment of the labile concentrations of metals in porewaters. Laboratory measurement of ROL in different species of Typha has also been successfully carried out by Matsui Inoue and Tsuchiya (2008). Sediment porewater includes the

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most labile contaminant fraction; assessment of metals in porewater is an effective means of determining the risk to receptors (Moreno-Jiménez 2011). Rhizon sediment porewater samplers offer an effective and non-invasive means to extract sediment porewater in-situ from the root zone of wetland plants over an indefinite period. Rhizon samplers have been successfully used to measure trace element concentration and mobility under field conditions (Clemente et al. 2008).

The aim of the study was to assess the photosynthetic performance of the wetland plants – B. articulata, C. appressa, S. validus (Cyperaceae), and P. australis

(Poaceae) growing in wetland sediment impacted by saline mine leachate characterised by excess concentrations of Ca, Cu, Mg, Mn, Na, S, and Zn. In addition, the impact of these plants on porewater pH, ORP and the concentrations of Ca, Cu, Mg, Mn, Na, S , Zn, and mineral N [NH4- N, NOx -N (NO2-N, NO3-N)]

(hereafter referred as Mineral N) was also measured. The fresh and dried biomass of each species and accumulation of Ca, Cu, Mg, Mn, Na, S, and Zn in roots and shoots along with translocation parameters including bioconcentration factor

(BCF), and translocation factor (TF) was also determined.

4.2 Statistical analysis

Concentrations of Ca, Cu, Mg, Mn, Na, S, and Zn in porewater were subjected to a two-way ANOVA analysis. Bonferroni test was used to generate letters for means.

All data were analysed using GenStat 15th edn. (2012). Porewater concentrations of

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each element are presented as means and standard deviations of control and treatment of each species.

Concentrations of Ca, Cu, Mg, Mn, Na, S, and Zn in porewater were subjected to a repeated measure (REML) analysis. Concentrations of Mn, mineral N and ORP

were log10 transformed. Bonferroni test was used to generate letters for means. All data were analysed using GenStat 15th edn. (2012). Significance value was set at

(P<0.05).

Gas exchange data was subjected to a repeated measure (REML) analysis.

Bonferroni test was used to determine significant differences between individual means.

All data were analysed using GenStat 15th edn. (2012). Significance value was set at (P<0.05).

The relationships of Ca, Cu, Mg, Mn, Na, S, and Zn in porewater were correlated

-2 -1 with photosynthetic performance (CO2m s ), non-photo chemical quenching

(NPQ), and fresh weight (grams) to determine significant relationships.

Correlations were selected for significance; significant correlations were then subjected to regression analysis. Significance value was set at (P<0.05). All data were analysed using GenStat 16th edn. (2013)

4.3 Material and methods

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4.3.1 Plant materials and treatment

A greenhouse experiment was conducted in a temperature controlled glasshouse

(30-15/15-10°C day/night temperatures, 50-70% relative humidity), with natural ambient light conditions at Charles Sturt University, Australia. Seedlings of S. validus, C. appressa, P. australis, and B. articulata were obtained from a local nursery at approximately 6-months old. The seedlings were grown in 5 L pots containing wetland sediment collected from a waste-rock leachate storage pond

(treatment) receiving saline leachate and a nearby (approx. 500 m) pond (control) not receiving saline leachate. In total 64 plants were assessed, with 8 control and 8 treatment for each species. The top 20 cm of wetland sediment was randomly collected from an area of approximately 50 m2 from each site, care was taken to ensure minimal disturbance of the sediment when placing in the pots. The pots were returned to the glasshouse immediately following collection.

The plants growing in sediment from the leachate pond were regularly supplied with leachate collected from the leachate pond at the mine site. Plants growing in the sediment from the control site were regularly supplied with leachate collected from the control pond. Throughout the experiment leachate and water was collected from the mine site as required, approximately 60 L was used each week. The pots were inspected daily and were watered when required to maintain a waterlogged state to ensure anaerobic conditions in the pots. The chemical composition of leachate and control pond water is presented in Table 14. The leachate has a mean

2- pH of 7.9, with elevated levels of Ca, Cu, Mg, and SO4 , and as such is determined 111

as saline leachate (Plumlee et al. 1999), concentrations of elements were determined as per porewater described below. The experiment was run for eight months from March 2012 – October 2012.

Table 14. Chemical composition (mg/L) of mine waste-rock dump leachate and control pond water collected during study.

Element Leachate pond Control pond Ca 502.8 495.4 Cu 0.169 0.013 Mg 423.8 357.5 Mn 0.617 0.37 SO4 2872 820.6 Zn 0.125 0.002

4.3.2 Particle size assessment

Particle-size analysis was performed on the wetland sediment to determine the particle-size distribution. Air-dried samples were added to 200 mL of deionised water and sodium hexametaphosphate and shaken on an end-over-end shaker for 16 hours at 15 rpm. Following shaking, hydrometer readings were recorded from the sediment samples in 1 L measuring cylinders at 5, 30, 93 and 420 minutes. The summation percentage (P) was calculated along with the particle-size diameter (D; in mm) and the percentage for each fraction was determined by plotting P against D

(Standards Association of Australia 1976).

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Table 15. Particle size analysis of sediment collected from control and treatment pond.

Fraction (mm) Control pond (%) Treatment pond (%) clay = <0.002 mm 15.16 16.13 silt = 0.002–0.02 mm 26.8 37.44 fine sand = 0.02–2.0 mm 54.8 42.35 coarse sand = 0.02–2.0 mm 3 3.35 gravel = >2.0 mm 0.39 1.3

4.3.3 Mine waste-rock dump leachate

The trialled plants growing in the sediment from the leachate pond were watered regularly with leachate and correspondingly water from the control pond was used for plants growing in the control sediment. Throughout the experiment, sediment remained waterlogged.

4.4.1 Carbon assimilation

4.4.1.1 Gas exchange measurements

Leaves of each species and the same physiological age were selected for gas- exchange measurements at two-monthly intervals from April to October. Net- photosynthetic rate (Pn), stomatal conductance (gs), internal CO2 concentration

(Ci), and transpiration (T) were assessed on intact leaves using a portable photosynthesis system (LI-COR 6400XT , Lincoln, Nebraska), equipped with a clamp-on leaf cuvette that exposed 2 cm2 of leaf area. Light was maintained at 500

-2 -1 µmol m s for all measurements, and humidity at 50 % ± 2 %, CO2 was

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–2 –1 maintained at a constant level of 380 µmol m s using a LI-6400-01 CO2 injector

(Lincoln, Nebraska) with a high pressure liquefied CO2 cartridge source.

4.4.1.2 Chlorophyll fluorescence measurements

Leaf-fluorescence measurements were conducted on leaves as per gas exchange measurements and determined using a portable-photosynthesis system (LI-COR

6400XT, Lincoln, Nebraska) equipped with a 6400-40 Leaf Chamber Flurometer

(LCF). Leaves were acclimated to environmental light to determine the efficiency of photosystem II (PHiPS2) (Genty et al. 1989), followed by non-photochemical quenching (NPQ), leaves were dark adapted with dark adapting clips (LI-COR

9964–091, Lincoln Nebraska) for 2 h before determination quantum efficiency of

Fv/Fm (Fo the fluorescence level of a dark-adapted plant with all PSII primary acceptors ‘open’[QA fully oxidized]. Fm is the maximal fluorescence level achieved upon application of a saturating flash of light, such that all primary acceptors

‘close’ (QA fully reduced). Variable fluorescence, Fv, is the difference between F0 and Fm.

4.3.1.3 Leaf tissue preparations for carbon and nitrogen isotope ratio measurements

Populations of B. articulata, C. appressa, P. australis, and S. validus were grown for eight months (March 2012 – October 2012). At the conclusion of this period

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plants were harvested and dried in an oven at 70°C for 48 hours before isotopic analysis. Using a plant grinder with 0.5 mm mesh screen (J.P. van Gelder & Co.

Pty Limited, Woy Woy, Australia), the dried leaf materials were ground to a fine powder.

4.4.1.4 Carbon and nitrogen concentration determinations and isotope ratio measurements

Three sub samples of the above leaf powder from the control and treated samples of each species were analysed at the Stable Isotopes for Biosphere Science laboratory,

Texas A&M University, College Station, TX 77843, USA for their carbon (δ13C,) and nitrogen (δ15N) isotope-ratio measurements, as well as their C and N concentrations. The δ13C, δ15N, and C and N concentrations were determined using an elemental analyser (Carlo Reba EA-1108, CE Elantech, Lakewood, NJ) interfaced with an isotope ratio mass spectrometer (Delta Plus, Thermo Electron

Corp., Waltham, MA) operating in continuous flow mode. Carbon isotope ratios

3 are presented in δ notation: δ = [(RSAMPLE – RSTD) / RSTD] x 10 where RSAMPLE is the

13 12 13 12 C/ C ratio of the sample and RSTD is the C/ C ratio of the V-PDB standard

(Coplen 1996). Precision of duplicate measurements was ± 0.1 ‰. N isotope ratios

3 are also presented in δ notation: δ = [(RSAMPLE – RSTD) / RSTD] x 10 where RSAMPLE is

15 14 15 14 the N/ N ratio of the sample and RSTD is the N/ N ratio of the atmospheric dinitrogen (Mariotti 1983). Precision of duplicate measurements for δ15N was < 0.2

‰.

4.4.1.5 Measurement of Rubisco Levels

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4.4.1.6 Preparation of leaf-tissue extracts

Fresh leaf material from B. articulata, C. appressa, P. australis, and S. validus was harvested at the conclusion of the experiment and immediately stored at –80° C before transporting them overnight to the University of Nebraska, Lincoln. A 100 mg of sample plant material was homogenised in a chilled mortar with a pestle in a

25 mM cold Bicine buffer (pH 7.8) containing 1 mM EDTA (ethylene diamine tertracetic acid) and 2% PVP (polyvinylpyrrolidone) and washed sea sand. Extract was centrifuged at 15,000 rpm (27,000 x g) for 15min at 4°C. After removing the pellet, the supernatant was transferred to another tube and its volume measured.

4.4.1.7 Determination of Total Soluble Protein:

Total soluble protein in the leaf extracts was determined by the Bradford method

(Bradford, 1976) using the Bio-Rad protein assay reagent and BSA (Bovine Serum

Albumin) as the standard.

4.4.1.8 Rubisco quantitation

Amounts of Rubisco in the leaf extracts were determined by SDS – PAGE gel electrophoresis (BioRad, Mini Protean® 3 cell; BioRad, CA, USA) using a Rubisco

Quantitation Kit (Agrisera, Vännäs, Sweden). An identical amount of protein from each species’ tissue leaf extract was loaded in to the wells of two Criterion precast

116

gels (Bio-rad) along with pre-stained molecular weight standards. Electrophoresis was carried out for one hour at 100 volts constant power. One gel was stained with

Coomassie Blue R-250 for an hour and later destained in a solution containing 30% methanol and 10% acetic acid, overnight. The gel was rinsed with water and saved.

The other Gel was prepared for immunoblotting by following the standard protocols (Towbin et al., 1979). The proteins from the gel were blotted onto a nitrocellulose membrane using the Criterion Blotting system from Bio-Rad. Blots were blocked in blotto (5% non fat dry milk in PBS, phosphate buffered saline).

The blot was then probed with purified Rubisco large subunit antibodies from the

Rubisco quantitation kit (Agrisera) followed by the detection using the Super

Signal West Pico chemiluminescent substrate kit (SuperSignal®West Pico,

ThermoScientific, Rockford, IL, U.S.A) according to the manufacturer’s instructions. The x-ray image was scanned and read using TotalLab TL100 v2006b, 1D gel analysis (TotalLab Ltd, Keel House, Garth Heads, Newcastle upon

Tyne).

4.4.1.9 Rubisco Standard Curve:

A standard curve using the purified Rubisco from the Rubisco quantitation kit was first constructed by following the same procedures as described for the leaf extract electrophoresis except for different amounts of Rubisco loaded into the wells.

4.5 Mineral accumulation analysis

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All samples for mineral accumulation analysis were dried to constant mass at 70°C for 48 h. Dried plant samples were then pulverized in a plant grinder with 0.5 mm mesh screen (J.P. van Gelder & Co. Pty Limited, Woy Woy, Australia).

Subsamples (0.5 g) were digested with 5 mL HNO3 at 60°C for 18 h, after cooling the suspension was adjusted to 50 mL with deionised water. Concentrations of Ca,

Cu, Mn, Mg, S, and Zn in plant roots and shoots were analysed by ICP–OES (710

ES, Varian, California, USA) at the Environmental and Analytical Laboratory

(EAL), Wagga Wagga, NSW, Australia. Each sample was digested as three replicates. Accuracy and precision of the digestion procedure and analysis was verified with reference material (CRM024–050, RTC, Sigma Aldrich, St Louis,

U.S.).

4.6 Porewater collection

At the conclusion of the experiment and harvesting of plant material porewater was extracted from each pot to determine the bioavailable fraction of Ca, Cu, Mg, Mn,

Na, S, and Zn. Ten-cm Rhizon porewater samplers (19.21.01 Eijkelkamp

Agrisearch Equipment, Giesbeek, The Netherlands) were inserted into the side of each pot. A Luer-lock syringe was attached to each sampler for extracting the porewater. The first 10 mL was discarded and the remainder was collected to ensure that fresh porewater was extracted from the immediate vicinity of roots. The concentrations of Ca, Cu, Mg, Mn, Na, S, and Zn in porewater are presented as means of eight controls and eight treatment of each species.

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4.7 Growth measurements

At the conclusion of the experiment fresh and dried mass (DM) were determined.

Dry mass were obtained by desiccating the roots and shoots at 70°C till they achieved a constant mass. Bioconcentration factor (BCF) was used as a measure of the uptake of Ca, Cu, Mn, Mg, S, and Zn from the treatment solution: BCF = element concentrations in tissues (mg/kg-1)/element concentration in solution

(mg/L-1). The translocation factor (TF) was calculated as a measure of ions Ca, Cu,

Mn, Mg, S, and Zn transport from roots to shoots: TF% = (ion content in shoot/ion content in root) X 100.

4.8 Results

4.8.1 Carbon assimilation

4.8.1.1 Gas exchange measurements

-2 -1 Photosynthetic gas exchange (Pn µmol CO2m s ) between control and treatment of each species was not significantly (P<0.05) different, although a significant difference occurred between each species (Figure 10). The greatest gas exchange occurred in the treatment populations of P. australis at 13.4 compared with the

-2 -1 control at 10.6 (CO2m s ). The photosynthetic gas exchange of the remainder of the control populations of each species were less than the treatment populations.

-2 -1 Stomatal conductance (gs mol H2O s ) did not differ significantly between control and treatment populations of each species. Internal CO2concentration (Ci µmol

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mol-1) did not differ significantly between control and treatment of each species, although control populations of C. appressa were significantly less than treatment populations of B. articulata, and P. australis, and control and treatment populations

-2 -1 of S. validus. Transpiration (T mol H2Om s ) did not differ significantly between treatment and control of each species, the greatest rate of transpiration occurred in control populations at 4.29.

Quantum efficiency of PSII (Fv/Fm) between control and treatment of each species, and within all tested species was not significantly inhibited, values of Fv/Fm ranged from as low as .768 in treatment populations of B. articulata to .817 in treatment populations of S. validus (Figure 11). A significant (P<0.05) difference of PSII efficiency (PhiPS2) occurred between control (.437) and treatment (.515) populations of S. validus only. Control (.472) and treatment (.526) populations of

P. australis differed significantly from control (.350) and treatment (.339) populations of B. articulata. Values of NPQ differed significantly between control

(1.18) and treatment (1.56) of B. articulata, the NPQ values of treatment populations of B. articulata were the highest compared control and treatment populations of the remainder of the species. The lowest NPQ values occurred in P. australis with .55 (control) and .37 (treatment), these values being significantly less than all other tested species.

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(a) (b) (c)

c c d

bc bc bc bc

ab ab

bcd a

abcd

abc

µmol mol µmol

mol H mol

a ab

g C

n µmol CO µmol n

(d)

bcd

abc

bcd

T mol T H mol

121

(a) (b) (c)

d e

ab cd

a ab

bc de

ab cd

ab a

NPQ ab

PhiPS2

122

4.8.1.2 Carbon and nitrogen leaf percentages

Carbon percentages did not differ significantly between control and treatment populations of each species or between each species (Table 15). Nitrogen percentages were significantly (P<0.05) greater in treatment populations of all species compared with control populations with the highest percentage of nitrogen recorded in treatment populations of P. australis.

C % N % Control Treatment Control Treatment B. articulata 42.59a ± 1.59 44.84a ± 1.52 0.957a ± 0.04 1.417b ± 0.06 C. appressa 44.90a ± 0.82 43.48a ± 5.31 0.817a ± 0.02 1.637bc ± 0.16 P. australis 41.14a ± 0.45 41.13a ± 2.97 1.687bc ± 0.05 2.943d ± 0.17 S. validus 37.60a ± 1.33 38.86a ± 2.56 1.037a ± 0.03 1.833c ± 0.14

4.8.1.3 Carbon and Nitrogen isotope ratios

Carbon isotope ratios (δ13C) differed significantly (P<0.05) between control populations of B. articulata, C. appressa, and S. validus (Table 16). The least negative value occurred in control populations of P. australis with -30.23, treatment populations of P. australis were -29.98. Nitrogen isotope Rations (δ15N) differed significantly between control and treatment populations of each species, treatment populations of C. appressa and P. australis also differed significantly.

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δ13C δ15N Control Treatment Control Treatment B. articulata -28.76b ± 0.015 -27.27c ± 0.11 6.37b ± 0.05 17.38e ± 0.39 C. appressa -28.47b ± 0.12 -27.00c ± 0.19 7.61c ± 0.22 23.84f ± 0.15 P. australis -30.23a ± 0.08 -29.98a ± 0.06 8.02c ± 0.24 15.45d ± 0.28 S. validus -30.11a ± 0.06 -28.72c ± 0.005 4.95a ± 0.15 16.73e ± 0.28

4.8.1.4 Rubisco quantitation

Amounts of Rubisco were determined through gel electrophoresis and Rubisco standards were used for making a standard curve. For all species the treatment populations contained a greater quantity of Rubisco compared with control populations (Figure 12). The greatest difference between treatment and control occurred in S. validus.

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Table 18. Rubisco levels (nmol) per mg of total soluble leaf protein in B. articulata, C. appressa, P. australis, and S. validus.

Control Treatment B. articulata 18.5 19.4 C. appressa 24.6 24.9 P. australis 25.7 30.4 S. validus 9.8 39.1

C T C T C T C T (a) X C X C X C X C X X X X

B. articulata C. appressa P. australis S. validus (b)

2.25 1.80 1.50

1.20 Figure 12. Immunoblot showing.90 the different amounts .60 of Rubisco (large subunit) in the different tissue extracts of B. articulata .3, C.(pmol) appressa , P. australis, and S. validus when identical amounts of soluble protein were loaded on the gel; C = control and T = treatment for each species (a), immunoblot showing Rubisco used for constructing the standard curve (b).

4.8.1.5 Gas exchange relationships

-2 -1 Photosynthetic gas exchange (Pn µmol CO2m s ) of all species declined significantly as concentrations of Mg, Na, and S increased in porewater. Non-photo chemical quenching (NPQ) values also declined significantly as concentrations of

Mg, and Na increased in porewater; and freshweight also declined significantly as concentrations of Cu, Mn, and Zn increased in porewater.

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(a) (b)

1 1

- -

s s

2 2

- -

m m

2 2

n µmolCO n µmolCO

P P

Mg mg/L Na mg/L

NPQ

n µmolCO

S mg/L Mg mg/L

(e) (f)

NPQ M/grams

Na mg/L Cu mg/L 126

(g) (h)

rams

M/grams M/g

Mn mg/L Zn mg/L . . . ▲ B. articulata, ▼...... C. appressa, ■- - - P. australis, ♦----S. validus

4.9 Mineral accumulation analysis

Mean accumulation of Ca and Cu was greatest in treatment roots of C. appressa,

mean accumulation of Mg was greatest in shoots of treatment populations of C.

appressa, mean accumulation of Mn was greatest in control shoots of S. validus,

mean accumulation of Na was greatest in treatment roots of C. appressa, mean

accumulation of S was greatest in treatment shoots of C. appressa, and mean

accumulation of Zn was greatest in treatment roots of C. appressa (Table

22).Translocation from roots to shoots of Ca, Cu, Na, and Zn was minimised by C.

appressa, and the greatest proportion of accumulation was recorded in the roots of

the treatment plants, while the greatest accumulation of Mg occurred in the shoots

of the treatment plants. Translocation by B. articulata was restricted to roots for

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Cu, Mg, and S in both control and treatment plants; while significantly more Na was translocated to the shoots of both treatment and control plants. Root to shoot translocation for P. australis showed a greater level of accumulation of Ca, Mg,

Mn, S, and Zn in shoots of both control and treatment plants; translocation by S. validus showed a greater proportion of Ca, Mg, Mn, and Zn in shoots of treatment and control plants.

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B. articulata C. appressa P. australis S. validus Ca Ca Ca Ca Shoots 2903e ± 45.9 2865e ± 111.5 5215h ± 365.7 5533h ± 300.2 2815e ± 96.05 2298cd ± 96.7 5082h ± 35.1 4470g ± 177.9 Roots 2658de ± 133.3 5075h ± 264.9 3740f ± 71.5 7015i ± 179.2 1087a ± 44.2 1943bc ± 109.8 1483ab ± 81.2 1355a ± 616.5 Cu Cu Cu Cu Shoots 13.1a ± 0.2 18.9a ± 1.36 16.8a ± 0.44 43.9ab ± 2.07 14.8a ± 3.55 32.8ab ± 2.63 15.5a ± 2.10 45.9ab ± 0.72 Roots 23.5a ± 0.7 70.3ab ± 1.78 29.1ab ± 3.36 263c ± 1.14 17.8a ± 3.52 203c ± 11.4 30.4ab ± 1.13 98.5b ± 3.30 Mg Mg Mg Mg Shoots 1340ab ± 19 2250d ± 62 3637f ± 57.9 6188j ± 57.7 1737c ± 56.5 2858e ± 114.1 3988g ± 487.5 4668h ± 203 Roots 1597bc ± 38.3 2482d ± 98.7 2353d ± 77.6 5160i ± 90.8 1273a ± 25.8 2498d ± 154.5 1430abc ± 49.4 1597bc ± 76.1 Mn Mn Mn Mn Shoots 355bc ± 3.54 93a ± 1.81 416c ± 181 191a ± 5.8 500cd ± 23.06 141a ± 4.90 1927f ± 220 232ab ± 9.42 Roots 176a ± 5.8 143a ± 4.46 735c ± 31.9 613de ± 27.9 175a ± 3.82 160a ± 9.70 235ab ± 10.07 173a ± 6.77 Na Na Na Na Shoots 2755e ± 33.91 5577i ± 117.8 4608h ± 127.81 5968j ± 51.15 865bc ± 29.84 972c ± 32.34 539a ± 29.83 497a ± 32.34 Roots 2487e ± 78.66 3755g ± 140.97 3400f ± 119.5 6237j ± 126.44 987c ± 19.54 1582d ± 158.17 967c ± 19.54 554ab ± 158.17 S S S S Shoots 2478c ± 38.17 3585e ± 112.56 6038h ± 139.77 9432j ± 225.6 3470e ± 86.49 5087f ± 219.33 823b ± 105.33 708b ± 29.16 Roots 2865d ± 74.8 5595g ± 203 5175f ± 103.68 7132i ± 201.44 2703cd ± 51.64 3508e ± 223.73 176a ± 6.05 263a ± 7.52 Zn Zn Zn Zn Shoots 12.23a ± 0.16 32.4f ± 1.68 15.25ab ± 0.4 26.57e ± 1.05 21.93de ± 0.9 176.17k ± 6.7 21.27cd ± 1.53 43.08h ± 1.6 Roots 16.75abc ± 1.09 40.12h ± 1.37 34.82fg ± 0.96 116.5j ± 3.02 19.02bcd ± 0.29 103.27i ± 4.43 20.07bcd ± 0.92 38.28gh ± 1.61 Control Treatment Control Treatment Control Treatment Control Treatment

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Figure 14. Image showing pots with Rhizon porewater samplers inserted into the side of each pot.

4.9.1 Plant growth measurements

Mean fresh mass of B. articulata, C. appressa, and S. validus were significantly

(p<0.05) reduced when raised in the treatment conditions; while mean fresh mass of P. australis raised in treatment conditions outperformed the control. Mean dry mass of C. appressa and S. validus were significantly (p<0.05) reduced; while mean dry mass of B. articulata and P. australis were not significantly reduced

(Figure 13).

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b (a) (b) d bcd cd

abcd ab abc ab ab ab a a a a a

a M/grams

Figure 155. Fresh (a) and dry (b) mass of B. articulata (B.a), C. appressa (C.a), P. australis (P.a), and S. validus (S.v). Mean data presented with standard deviation, columns with different letters indicate significant difference between control (white columns) and treatment (black columns) at l.s.d at 5 % level, n=6.

4.9.2 Bioconcentration and translocation

Mean BCF of each element differed significantly between each species grown in treatment and control conditions [Table 23 (a)]. For B. articulata BCF of Cu, S and

Zn were significantly greater in control populations, and Na was significantly greater in treatment populations; for C. appressa Cu, Mg, S and Zn were significantly greater in control populations, while Ca was significantly greater in treatment populations; for P. australis BCF of Cu, Mg, Na, and S were significantly greater in control populations, and Mn and Zn were significantly greater in treatment populations; for S. validus significantly greater BCF values of

Cu, Mg, Mn, Na, S and Zn occurred in control populations.

Mean TF also differed significantly between treatment and control of each species

[Table 23 (b)]. The TF for Ca and Mn was significantly greater in control populations of B. articulata compared with treatment populations; TF of Ca, Mg, and Zn was significantly greater in treatment populations compared with treatment

131

populations; for P. australis the TF of Ca, Cu, and Mn were significantly greater in control populations, while the TF of Zn was significantly greater in treatment populations; the TF of Mn and S was significantly greater in control populations of

S. validus, while the TF of Na was significantly greater in treatment populations.

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(a) B. articulata C. appressa P. australis S. validus Ca Ca Ca Ca 5.37ab ± 0.08 6.36b ± 0.24 9.08c ± 0.31 12.64e ± 0.68 4.79a ± 0.18 5.45ab ± 0.22 11.36d ± 1.37 10.74d ± 0.42 Cu Cu Cu Cu 466.7bc ± 7.02 196.2a ± 14.1 578.2de ± 15.2 369.5b ± 17.4 482.9cd ± 115.5 232.2a ± 18.6 606.5e ± 82.3 177.5a ± 6.85 Mg Mg Mg Mg 2.06a ± 0.02 1.73a ± 0.04 5.29e ± 0.08 4.35d ± 0.04 3.64c ± 0.11 1.76a ± 0.07 6.23f ± 0.76 2.93b ± 0.12 Mn Mn Mn Mn 41.7ab ± 0.41 54.2b ± 1.05 30.5ab ± 0.66 34.6ab ± 1.04 21a ± 0.96 90.2c ± 3.10 298.2e ± 34.1 233.8d ± 9.51 Na Na Na Na 7.1e ± 0.08 11.84g ± 0.25 11.32f ± 0.31 11.61fg ± 0.09 2.72d ± 0.09 1.69c ± 0.05 1.3b ± 0.16 0.89a ± 0.03 S S S S 1.71d ± 0.02 1.44c ± 0.04 3.82g ± 0.08 3.46f ± 0.08 2.78e ± 0.06 1.63d ± 0.07 0.55b ± 0.07 0.23a ± 0.009 Zn Zn Zn Zn 2597h ± 34.7 272b ± 14.1 2472g ± 62.5 120a ± 4.75 832d ± 34.1 1218e ± 46.1 1956f ± 140.2 477c ± 17.6 Control Treatment Control Treatment Control Treatment Control Treatment

(b) B. articulata C. appressa P. australis S. validus Ca Ca Ca Ca 109.5bc ± 6.76 56.7a ± 4.96 139.5c ± 6.05 79.2ab ± 7.36 258.9d ± 7.82 118.4c ± 6.3 342.9e ± 41.3 330e ± 13.1 Cu Cu Cu Cu 55.65ab ± 2.25 26.85a ± 1.63 58.35ab ± 7.45 35.02a ± 50.9 88.20b ± 33.6 16.13a ± 1.31 50.92ab ± 7.29 18.86a ± 1.14 Mg Mg Mg Mg 84a ± 2.44 90.8ab ± 5.2 154.6d ± 4.77 120bc ± 2.54 136.4cd ± 5.42 114.8bc ± 9.63 279.5e ± 37.8 293e ± 20.4 Mn Mn Mn Mn 133

202.4c ± 7.07 65.2ab ± 1.71 66.5ab ± 3.05 31.2a ± 2.12 286.1d ± 11.4 88.5ab ± 5.07 822.3e ± 112.5 134.3bc ± 10.5 Na Na Na Na 110.9ab ± 3.84 148.7ab ± 7.57 135.7ab ± 5.79 95.7a ± 1.4 87.7a ± 3.66 62.1a ± 8 55.8a ± 7.96 251.7b ± 222.2 S S S S 86.6ab ± 3.3 64.2a ± 4.29 116.7bc ± 2.28 132.3bc ± 5.12 128.4bc ± 3.59 145.6c ± 13.6 470.3e ± 69.7 269.7d ± 11.1 Zn Zn Zn Zn 73.26c ± 4.29 80.92c ± 6.34 43.83b ± 1.63 22.82a ± 1.06 115.36d ± 5.02 170.84e ± 9.55 106.29d ± 10.2 112.69d ± 6.04 Control Treatment Control Treatment Control Treatment Control Treatment

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4.9.3 Elements in porewater

Calcium: Mean Ca concentrations (Figure 15 a) increased significantly (P<0.05) in the control over the sampling period replace with days in C. appressa (340.9,

495.9, 571.1 mg/L) and P. australis (384.8, 540.8, 585.8 mg/L). The final mean concentrations of Ca did not differ significantly between control and treatment of each species with the exception of P. australis (control 585.5 mg/L, treatment

421.3 mg/L).

Copper: Mean Cu (Figure 15. b) concentrations significantly increased during the sampling period in treatment populations of P. australis only (0.051, 0.093, 0.141 mg/L). Samples collected at the first sampling period did not differ significantly between treatment and control of each species; although samples of each species obtained in the final collection differed significantly between control and treatment as follows: B. articulata (control 0.028 mg/L, treatment 0.0962 mg/L), C. appressa

(control 0.0288 mg/L, treatment 0.118 mg/L), P. australis (control 0.031 mg/L, treatment 0.141 mg/L), S. validus (control 0.025 mg/L, treatment 0.104 mg/L).

Magnesium: Mean Mg (Figure 15. c) concentrations significantly increased over the sampling period in treatment populations of all tested species, B. articulata

(583, 955.8, 1276 mg/L), C. appressa (513.9, 1010.5, 1422 mg/L), P. australis

(558.9, 1108.9, 1620.6 mg/L), S. validus (624.2, 1176.6, 1515.2 mg/L). As in the case of Cu, the mean concentrations of Mg did not differ significantly between control and treatment at the first sampling point, but at the completion of the 135

experiment the control and treatment differed significantly, B. articulata (control

662.4 mg/L, treatment 1276 mg/L), C. appressa (control 694.7 mg/L, treatment

1422 mg/L). P. australis (control 484.1 mg/L, treatment 1620.6 mg/L) S. validus

(control 639.5 mg/L, treatment, 1515.2 mg/L).

Manganese: Mean Mn (Figure 15. d) concentrations did not significantly increase in control or treatment populations of each species. At the conclusion of the experiment mean concentrations of Mn in treatment populations of P. australis

(control 1.266 mg/L, treatment -0.756 mg/L) were significantly less than control populations.

Sodium: Mean Na (Figure 15. e) concentrations increased significantly between the first and final sampling period in both treatment and control of all species, although the mean concentrations of Na at the completion of the experiment did not differ significantly between control and treatment populations of B. articulata and C. appressa, while the final mean concentration did differ significantly between control and treatment of P. australis (control 324.4 mg/L, treatment 573.9 mg/L) and S. validus (control 412.7 mg/L, treatment 556.3 mg/L).

Sulphur: Mean S (Figure 15 f) concentrations increased significantly between the first sampling period and the final in treatment populations of C. appressa (1554,

1965, 2718 mg/L), P. australis (1497, 2096, 3111 mg/L) and S. validus (1622,

2181, 2998 mg/L). In addition, the final mean concentration of each species

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differed significantly between control and treatment of each species, B. articulata

(control 1458 mg/L, treatment 2480 mg/L), C. appressa (control 1558 mg/L, treatment 2718 mg/L), P. australis (control 1267 mg/L, treatment 3111 mg/L), S. validus (control 1495 mg/L, treatment 2998 mg/L).

Zinc: Mean Zn (Figure 15 g) concentrations did not differ significantly between the first and final sampling period in treatment and control populations of each species, with the exception of C. appressa. At the completion of the experiment the mean concentration of Zn for C. appressa was 0.008 mg/L for control and 0.221 mg/L for treatment. For C. appressa, P. australis and S. validus a significant increase occurred in treatment populations in the second sampling period, but declined in the final sampling period. Zinc was the only element to display this increasing and decreasing trend from the first to third sampling period.

Mineral N: Mean Mineral N (Figure 15 h) concentrations decreased significantly in control populations of C. appressa (1.268, –0.401, –0.061 mg/L), P. australis

(1.243, –0.303, –0.227 mg/L), and S. validus (1.28, 0.34, 0.302 mg/L). The final treatment Mineral N concentrations were significantly greater than the control of each species with B. articulata (control 0.323 mg/L, treatment 1.849 mg/L), C. appressa (control –0.061 mg/L, treatment 1.528 mg/L), P. australis (control 0.227 mg/L, treatment 2.023 mg/L), S. validus (control –0.302 mg/L, treatment 1.788 mg/L).

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ORP: Mean values of ORP (Figure 15. i) increased significantly in treatment

population of B. articulata (14.57, 51.40, 122.06 mV), P. australis (8.97, 47.40,

116.76 mV), and S. validus (–5.40, 23.86, 115.96 mV). Although at the completion

of the experiment mean ORP values were not significantly different between

control and treatment.

pH: Mean values of pH (Figure 15 j) only increased significantly in control

populations of S. validus (6.638 > 6.668 > 7.094 pH) from the first to final

measurement. Values of pH ranged from 6.638 in the control populations of S.

validus to 7.596 in treatment populations of B. articulata.

a (Ca)

cde cde

bcde bcde

–

abcd bcde

a abcde abcd

abcd

abcde abcde abcd

abcd abcd abcd ab abc

abc

ab ab

mg / L / mg

Ca Ca

b (Cu)

ghi

–

h i

– –

eghi

c e

efgh

–

mg / L / mg

–

e a

–

Cu Cu

abcd

abcd abcd

ab abc

abc ab a ab

a abcd

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C (Mg)

ghi

fhi

fghi

mg / L / mg

efgh

–

Mg Mg

–

abcd a abcd

–

abc

a abcd

abc abc

abc

a ab

abc

a a a

d (Mn)

cdef

–

b f e

–

def

def def

–

)

bcde a

def

–

– f

–

abcd a

–

a a a

abc

mg / L (log / L mg

Mn Mn

e (Na)

def

cde

cd cde

mg / L mg

ab ab

ab ab ab

Na Na

ab ab

a a a

a a

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f (S)

hijk

–

b c

f e f

e e

–

mg / L / mg

– – – –

– –

a a a a

S S a a

–

abcd a

e abc abcd

–

a a a

g (Zn)

cdef

cef

–

cdef

–

mg / L / mg

Zn Zn

–

abc a

abc abc ab abcd

abc

a a a

a a a a

h (Mineral N)

ef def

)

def def

def

def def

def

–

bcd b

bcde cde

bcde

abc

mg / L (Log L (Log / mg

a a

Mineral N N Mineral

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i (ORP)

d ef

–

cdef

)

–

a f

–

– f

–

bcdef a

–

mV (Log (Log mV

abcde

ab abc

–

ab a

ORP ORP

j (pH)

–

m m

–

– –

f f

– –

f – –

c c

j i

–

– b d

– –

b a

–

a a

– –

–

a a

abcd

abc

30 h (Temp) a

20 b

15 Temp 10 5 0 1 2 3

Sampling time Figure 166. Mean concentrations (mg/L) of porewater extracted from 4 L plastic pots containing control and treatment wetland sediment. Figure 15. a–(Ca), b–(Cu), c–(Mg), d–(Mn), e–(Na), f–(S), g–(Zn), h (mineral N), i (ORP), j (pH), h (Temp). B.a C = B. articulata (Control), B.a T = B. articulata (Treatment), C.a C = C. appressa (Control), C.a T = C. appressa (Treatment), P.a C = P. australis (Control), P.a T = P. australis (treatment), S.v C = S. validus (Control), S.v T = S. validus (treatment). S1, S2, and S3 indicate the three sampling periods over the 8 month experimental period. Columns with different letters (superscript) indicate a 141

significant difference, significance value was set at (P<0.05). Standard deviation indicated by rods within columns. Log to base 10 values of ORP, Mineral N, and Mn less than 0.1 will be negative, graphs shown on log scale, n=8.

4.8 Discussion

The values of all elements, mineral N, ORP, and pH in porewater were influenced by the selected wetland plants. Plants influenced either a significant increase or a significant decrease in the concentrations of elements in porewater. At the first sampling point concentrations of Ca, Cu, Mg, Na, and S in porewater were not significantly greater in treatment conditions compared with the control. The values of Ca, Cu, Mg, Na, and S, however significantly increased at the completion of the experiment in the treatment porewater samples compared with the control samples.

The oxidation of the root-zone by each species was evident by the increase in ORP, since the values changed from ‒57.79 to 122.06. Values of ORP in waterlogged soils are generally greater in vegetated soils compared with unvegetated soils

(McKee et al. 1988). The ORP values found in the present study aligned with those found in porewater impacted by T. latifolia and G. fluitans growing in acid-neutral pH waterlogged tailings, which ranged between 50 and 100 ORP (Wright and Otte

1999).

The mobilisation of elements or the lack of it is dependent on the physical, chemical, and biological conditions of the substrate, and therefore it can be difficult to predict which plant will be efficient in element storage (Jacob and Otte 2002).

Although, a significant contributor of ROL is the free oxygen generated during

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photosynthesis, diffusion from leaves towards roots occurs during the active growth period, which induces an increased level of ROL in the sediment (Caffrey and

Kemp 1991; Connell et al. 1999; Hupfer and Dollan 2003). Throughout the experimental period, the plants were maintained in an active-growth phase in controlled glasshouse conditions, and the porewater concentrations of elements continued to increase during the trial period. The highest values of Cu, Mg, Na, and

S in porewater however occurred in treatment populations of P. australis, in a photosynthetic assessment of all presently trialled species (see Chapter 5), P. australis assimilated a high volume of CO2. Significant differences of ROL have been recorded to occur in the wetland species such as; ROL by Typha angustifolia,

T. latifolia, and T. orientalis when tested in laboratory conditions (Matsui Inoue and Tsuchiya 2008). In waterlogged Zn–Pb mine tailings, T. latifolia increased soluble Zn, whereas Glyceria fluitans (Poaceae) showed the least effect (Wright and Otte 1999).

In the present study, Mn concentrations in the treatment populations dropped throughout in P. australis and S. validus, and were significantly lesser than the control populations. The effect of vegetation on porewater composition in a natural wetland receiving acid-mine drainage varied across seasons, P. australis and

Eriophorum angustifolium influenced significantly less concentrations of Mn in

Northern- Hemisphere winter and summer (Batty et al. 2006). For T. domingensis growing in Wetlands impacted by saline-mine leachate and a nearby Cadiangullong

Creek not directly affected by mining activity at the Cadia gold—copper mine

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operations in Orange (NSW, Australia), the total and DTPA (diethylene-triamine- pentaacetic acid) extracted Mn in sediment were greater in the winter compared with the summer, and root and shoot accumulation was greater in the summer compared with the winter (Southern Hemisphere) (Adams et al. 2013). Less concentrations of elements in wetland sediment occur during the warmer months when vegetation grows actively. Root exudates including free oxygen cause precipitation of Fe resulting in plaques around the roots of wetland plants; similarly oxide plaques of Mn have also been identified in the presence and absence of Fe plaques (St-Cyr and Crowder 1990). Although the exact cause for Fe and Mn plaques around roots is unclear presently, that a combination of ROL where oxygen reacts with reduced Fe and Mn and accumulates them as plaques on the root surface is one explanation (Batty 2003), coupled with the release of exudates that include enzymes and the microorganisms in the root-zone are the possible mediators

(Yamada and Ota 1958; Neubauer et al. 2002). Decreasing concentrations of Mn in treatment populations of P. australis and S. validus occurred during an active growing phase and therefore at a period with greater ROL.

Porewater was characterised by significantly greater concentrations of Cu and Mg in treatment pots of all species, and S in treatment pots of P. australis and S. validus (Table 21). These three elements were also elevated in the treatment leachate compared with the control water (Table 19). This study has shown that each species accumulated greater amounts of elements from treatment conditions containing greater concentrations of elements. In particular, P. australis is known

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to respond positively to increased concentrations of elements and shows a significant relationship between concentrations in sediment and those in plant organs (Bonanno 2011). The response of plants to the presence of pollutants such as metals can alter the plants’ ability to control the uptake of those pollutants, increasing the uptake and sometimes causing problems to the performance capability of the plant (Almeida et al. 2009). Although, these effects depend on the pollutant, plant species, and the concentrations and time of exposure to the pollutant (Davis et al. 2002). With the exception of Cu and Mn accumulation in shoots of S. validus, and Zn accumulation in shoots of P. australis, C. appressa accumulated the greatest quantity of each tested element in roots and shoots of treatment plants. Carex appressa has been used successfully in biofiltration systems, in particular for the removal of N and P (Bratieres et al. 2008; Zinger et al.

2013). However, trials using C. appressa for phytoremediation of mine leachate, in particular saline-mine leachate are limited. In the present study, C. appressa and B. articulata responded positively to the treatment conditions and accumulated significantly greater quantities of Na in roots and shoots compared with the other trialled species, and C. appressa translocated significantly greater quantities of S from roots to shoots at 9342 mg/kg; the nearest shoot accumulation value of S was the control population of C. appressa at 6038 mg/kg.

In the present study, C. appressa accumulated significantly greater concentrations of Mg, Mn, Na, and S compared with P. australis; however, the biomass of C. appressa was significantly reduced. The final biomass of plants after a treatment

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period of eight months is a proven indication of the tolerance of plants to metal treatment (Köhl and Lösch 2004); and biomass production is known to be substantially constrained by saline conditions (Hasegawa 2013). The present study shows that P. australis, and S. validus in particular are the most suitable and promising species for phytoremediation of the saline-mine leachate. The biomass of

B. articulata and C. appressa was significantly reduced in treatment conditions and therefore their suitability for phytoremediation of saline-mine leachate may be unsuitable. It is evident that individual traits of translocation of excess concentrations of elements from roots to shoots in B. articulata and C. appressa tends towards greater translocation of Na in particular compared with P. australis and S. validus. Fresh and dry mass of treatment populations of C. appressa were significantly less than the control (Figure 13), whereas the fresh and dry mass of the other trialled species were not significantly impacted in the treatment conditions, with the exception of the fresh mass of B. articulata. The regression analysis presented in Figure 13 shows the negative impact increasing concentrations of Cu, and Zn in porewater had on the fresh weight of each species.

Total biomass of Tamarix chinensis (Tamaricaceae) decreased by 63.3 % when exposed to 16 g kg-1 Na CI in soil (Wang et al. 2011). Nutrient removal in C. appressa is attributed to its dense and fine root patterns, although species of

Cyperaceae, such as C. appressa, are usually non-mycorrhizal (Miller et al. 1999); mycorrhizal associations restrict excess accumulation of elements (Greger 2004).

Colonisation of Coccoloba uvifera (Menispermaceae) by ectomycorrhizal (ECM) fungus contributed towards less Na and CI accumulation than non-ECM plants

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(Bandou et al. 2006). Accumulation and translocation of Na by P. australis and S. validus was significantly less than C. appressa and B. articulata, despite all species being exposed to similar (values did not differ significantly between species) concentrations of Na in porewater. This study shows that the physiological traits of

C. appressa and B. articulata enable greater quantities of Na and S accumulation.

Excess accumulation of Mn is known to inhibit CO2 assimilation in many plants including white birch (Betula platyphylla Suk.) (Kitao et al. 1997) and rice (Oryza sativa L.) (Lidon et al. 2004). Manganese-tolerant genotypes have evolved Mn- detoxifying strategies, including chelation in the cytosol and sequestration into vacuole and endoplasmic reticulum (González and Lynch 1999). A common protective trait in some plants is to also minimise uptake of Mn via root exudation of organic acid anions (Mora et al. 2009) and/or decreasing Mn translocation from roots to shoots (Vose and Randall 1962). Wetland plants in particular are known to store greater concentrations of metals in belowground tissues (Batty and Younger

2004). Those species known to be sensitive to excess Mn translocate excess concentrations from roots to shoots, while those tolerant to Mn-excess accumulate high quantities in roots (Mora et al. 2009). The present study however has shown that wetland plants of the same species varied significantly in translocation of Mn in particular; translocation of Mn to shoots was significantly greater in control populations compared with treatment populations (Table 20). In particular, control populations of S. validus translocated a mean of 1927mg/kg of Mn to shoots, without an apparent impact to photosynthetic activity; correlation analysis of CO2

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assimilation and concentrations of Mn in porewater did not reveal a significant impact to CO2 assimilation due to exposure of increasing concentration in porewater. Furthermore, biomass of those plants in increasing concentrations of Mn in porewater was significantly greater (Figure 13).

The inhibitory effect of Mn toxicity on photosynthesis has been reported to be due to induced modifications of Rubisco (Houtz et al. 1988). Analysis of Rubisco revealed greater concentrations in treatment populations compared with control populations, and consequently P. australis assimilated greater amounts of CO2.

Nitrogen content in leaves is known to have a strong supporting relationship with photosynthetic capacity (Field and Mooney 1986); this is largely due to the proportion of leaf N (up to 75 %) in the chloroplasts, and soluble proteins particularly Rubisco (Evans 1989). The low rate of catalysis and poor affinity for

CO2 of Rubisco is partly compensated for with high concentrations in leaves and therefore results in strong correlations between the quantity of Rubisco and the rate of light saturated photosynthesis (Woodrow and Berry 1988). Nitrogen deficiency is known to be a frequent abiotic stress; chloroplast proteins of N deficient Triticum aestivum and Phaseolus vulgaris were degraded under such conditions (Feller et al.

2008). The toxicity of salinity-alkalinity (NaCI, NaHCO3) is also known to impact negatively on photosynthesis of P. australis (Deng et al. 2011) and Maize (Zea mays) (Deng et al. 2010). The photosynthetic performance of P. australis growing in salinized wetlands receiving wastewater high in N reduced the toxicity and assisted photosynthesis (Deng et al. 2011). Percentage of N in leaf material in each

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species was significantly greater in treatment populations compared with control populations. The present study did not determine that gas exchange in the treatment populations of each species were significantly inhibited compared with the control populations (Figure 10 a,b,c,d). Percentage of N in leaf material in each species was significantly greater in treatment populations compared with control populations. The significantly greater amounts of N in leaf material is likely due to the wetland receiving wastewater high in N; therefore reducing the toxicity and assisting photosynthesis (Deng et al. 2011). While the analysis of CO2 assimilation between treatment and control did not reveal a significant impact on photosynthesis; regression analysis of concentrations of Na, Mg, and S in porewater and photosynthesis revealed a significant reduction in photosynthesis when increasing concentrations of these elements occurred in porewater. These findings indicate that P. australis in particular has an innate capacity to accumulate significantly greater quantities of N and as such photosynthesis and plant performance are greatly benefited.

For S. validus a greater concentration of Rubisco occurred in leaves compared with

P. australis, while S. validus had a significantly lesser quantity of leaf N compared with P. australis. Despite this, P. australis had significantly greater assimilation of

CO2. This reveals that despite a lesser quantity of Rubisco and a greater quantity of leaf N, P. australis assimilated a greater quantity of CO2 than S. validus despite having a greater concentration of Rubisco. The degenerative impact of Cd and salinity on Rubisco is well known (Chaffei et al. 2004). Rubisco content and

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activity were lowered by 10 % and 60 % respectively when found with 0.83 nmol

Cd (Pietrini et al. 2003). Although, information regarding the impact of other heavy metals (e.g., Al, Cr, Fe, Mn) is limited (Galmés et al. 2013). The physiological status of the free-floating aquatic plant Salvinia natans was inhibited when accumulating up to 9 mg g-1 Cu, and 4 mg g-1 Zn; for Rubisco, Cu declined from a control of 26.18 to 11.8 nmol, and Zn declined to 13.64 nmol (Dhir et al. 2011).

For the presently tested species, leaf accumulation ranged between 18.9 and 45.9 mg/kg and for Zn 26.57 and 176.17 in treatment populations without reducing the amount of Rubisco compared with control populations.

Non-photo chemical quenching (NPQ) in treatment populations of B. articulata was significantly (P<0.05) greater than control populations. Non-photo chemical quenching is a photoprotective mechanism for dissipation of excess excitation energy (Demmig-Adams and Adams 1992), whereby a portion of photons are released as thermal energy rather than used to drive photosynthesis (Shangguan et al. 2000), this leads to a downregulation of PSII and avoids over-reduction of primary electron acceptors (Genty et al. 1989). The treatment populations of B. articulata appeared to protect the photosynthetic apparatus through a dissipation of some excitation energy through an increase in NPQ, while PSII was not significantly impacted. Likewise, photodamage of Caragana korshinskii by excess concentrations of NaCI was avoided by dissipation of excitation energy which was evident by an increase in non-photochemical quenching, while maximum efficiency of PSII was not affected (Yan et al. 2012).

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Carbon Isotope Ratios (δ13C) were significantly less negative in treatment compared with control of each species, with the exception of P. australis. The δ13C

0 values for C3 plants typically falls within –27.1 ± 2.0 /00 (O’Leary 1988). Several environmental factors are known to influence the isotope fractionation in C3 plants

(Madhavan et al. 1991); including soil salinity (Flanagan and Jeffries 1989).

Stomatal diffusion and carboxylation are also known to reflect δ13C values

(Madhavan et al. 1991). Differences of δ13C values in leaves indicate a difference in stomatal regulation and photosynthesis (Walker and Sinclair 1992). Nitrogen

Isotope Ratios (δ15N) were significantly more positive in treatment populations of each species compared with control populations. Plant δ15N is known to reflect the variable δ15N of external sources and isotope fractions, which occur during the assimilation, transport and loss of N (Yun and Ro 2008). Robinson et al. (2000) found N starvation significantly affected whole-plant δ15N, and that the values always became more negative than the control. In the present study the treatment populations were grown in wetland sediment containing significantly greater amounts of porewater mineral N compared with the control. The more positive nitrogen isotope ratios of treated plants reflect not only the fractionation associated with the uptake and partitioning of acquired nitrogen but also the nitrogen isotope signature associated with the source nitrogen.

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Chapter 5: General conclusion

5.1 Summary of experimental aims

The aim of this thesis was to evaluate the suitability of T. domingensis, B. articulata, C. appressa, P. australis, and S. validus for phytoremediation of saline mine-leachate. This study consisted of determination of Ca, Cu, Mg, Mn, Na, S, and Zn within wetland substrate receiving saline mine-leachate, and root-zone interactions [Chapters 2, 4], measurement of carbon- assimilation capabilities (gas exchange, isotopic signatures, Rubisco activity) [Chapters 3, 4], and accumulation and translocation traits [Chapters 2, 3, 4]. These trials provide an assessment of the ability of each species to accumulate excess concentrations of elements with minimal impact of stress as evident by reduced biomass and/or photosynthetic performance. Each species was trialled in the existing conditions of a functioning mine-site, with no attempt to reduce the impact on the tested plants by way of soil additives or conditioners.

5.2 Plant evaluation

The novel aspect of this thesis was trialling Australian native wetland plants for phytoremediation of saline mine-leachate. With the exception of P. australis and T. domingensis, little is known about the potential phytoremediation suitability of B. articulata, C. appressa, and S. validus. The inclusion of lesser studied species with those such as P. australis and T. domingensis with a history of many experimental 152

trials provided a suitable means of comparison. The central hypothesis was that the selected species would be suitable candidates for phytoremediation of saline mine– leachate.

Three strategies enable plants to survive on contaminated and metalliferous soils; metal excluders—effectively prevent metal from entering their aerial parts, but can contain large amounts of metals in their roots; metal indicators—accumulate concentrations in their aerial parts aligned with soil concentrations (Memon and

Schröder 2009). The most effective plants for phytoremediation have been determined as hyperaccumulator species, characterised by substantial metal concentration in shoots (Wei et al. 2009). However, to date, no wetland species have been identified as hyperaccumulators and moreover an established trait of wetland plants is the tendency to store greater concentrations of elements in roots compared with shoots. This trait has the two–fold benefit of protecting the photosynthetic mechanism from increased levels of heavy metals, which are known to inhibit photosynthesis and retaining metals within the sediment substrate (Batty and Younger 2004).

The present study therefore identified that the traits of metal excluders— accumulating large concentrations in roots with minimal impact to photosynthetic performance as ideal traits of wetland plants used for phytoremediation. Wetland plants however, have been found to significantly influence the available concentrations of elements within the rhizosphere due to release of oxygen via

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roots (Batty et al. 2006). The rhizosphere interactions by the selected plants were also studied as part of the evaluation; this enabled an analysis of three key attributes of wetland plants and their suitability to phytoremediation of saline mine–leachate.

5.2 Concentration of elements in substrates and root-zone interactions

Assessments of T. domingensis examined its role in the removal of Cu, Mn, and Zn from the leachate pond receiving saline mine leachate, This species has naturally colonised the mine-site waterways over the last 10 years and as such appeared tolerant of the conditions. Concentrations of Cu, Mn, and Zn in the leachate were significantly reduced at the NLP during the summer, indicating the potential of the naturally colonised plant populations to contribute to immobilisation of metal- impacted waters and sediment.

Examination of porewater revealed that all measured parameters were influenced by the plants. The free oxygen generated during photosynthesis and diffused from leaves to roots during the active-growth phase is a significant contributor of ROL

(Hupfer and Dollan 2003). The findings in this work indicate that the species capable of greater CO2 assimilation influences a greater level of immobilisation of

Mn in porewater concentrations due to precipitation onto plant roots.

5.3 Photosynthetic performance

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The impact of photosynthetic stress on B. articulata was shown in its significantly greater non-photo chemical (NPQ) quenching in treatment populations. The percentage of N was significantly greater in treatment populations. Leaf N is a strongly correlating factor with the photosynthetic capacity because of its large proportion of incidence in chloroplasts (Field and Mooney 1986). Therefore, the greater concentrations of N in leaves supported greater quantities of Rubisco and subsequently greater CO2 assimilation.

The isotopic ratios of N were significantly more positive in treatment populations compared with control populations. Plant δ15N is a well-known indicator of the externally sourced N and reduced concentrations of N internally, as was evident in the control populations of the selected plants.

5.4 Accumulation and translocation of elements

The shoot accumulation traits of each species in the present study followed similar trends, with the exception of shoot accumulation of Na by B. articulata and C. appressa with significantly greater shoot accumulation than P. australis and S. validus. In addition, Mg and S accumulation in C. appressa shoots was also significantly greater than the remaining species.

Carex appressa accumulated significantly greater concentrations of Mg, Mn, Na, and S when compared with B. articulata, P. australis, and S. validus. Although, the fresh biomass of C. appressa and B. articulata were significantly reduced 155

compared with the other species, whereas the fresh and dry mass of the other trialled species were not significantly impacted in the treatment conditions; biomass is an indicator of the impact of saline conditions (Hasegawa 2013). Both accumulation and translocation of Na by P. australis and S. validus was significantly lesser than C. appressa and B. articulata.

5.5 Conclusions

Wetland phytoremediation is a sustainable alternative to chemical and engineering technology for the treatment of mine-waste leachate. While alternative treatment technologies are reliant on external power sources they generate secondary waste streams, phytoremediation harnesses solar energy to power remediation and stabilisation of contaminated sites. The ability of plants to withstand the conditions imposed by excess concentrations of elements is a key component of successful phytoremediation. Several key wetland species have been utilised within constructed and natural wetlands as a key component of successful treatment of mine-waste leachate from both operating and completed mine projects throughout the world. Phytoremediation is a term encompassing many subtypes of plant-based remediation approaches, including phytoextraction, phytostabilisation, and phytomining. The use of wetland environments to treat mine-leachate exiting mines has largely focused on capture and retention of the elements of concern within the wetland substrate. The anoxic and reducing conditions of wetland substrates and a tendency of many wetland plants to store greater amounts of elements in

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belowground tissues are favourable conditions for capture and retention of excessive concentrations of elements

Throughout this study it was determined that the leachate generated within the mine waste-rock dump was characteristic of saline mine-leachate; containing elevated concentrations of Ca, Mg, and S. While Cu, Mn, and Zn were also present it was determined that these metals were unlikely to cause a significant impact on plant performance based on the concentrations within the porewater. To date extensive research has been conducted on acid mine-leachate, this study however has shed light on the lesser known impact and subsequent performance of wetland plants receiving saline mine waters.

The assessments of photosynthetic performance presented in Figures 10 and 11 did not reveal a significant impact, with the exception of NPQ of C. appressa; although photosynthetic performance did vary significantly between species, the greatest rate of CO2 occurred in P. australis. Regression analysis of concentrations of Mg, Na, and S in porewater with CO2 assimilation; Mg in porewater with NPQ, and Cu, Mn, and Zn in porewater with fresh weight revealed photosynthetic performance and fresh weight were significantly reduced as the concentrations of these elements increased in porewater over the trial period (Figure 13). Wetland plants are known to release oxygen into the rhizosphere, contributing to the formation of Fe (hydr-

)oxide precipitates around the roots; therefore enabling adsorption and co- precipitation of other elements within the rooting zone, this can reduce

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concentrations of Fe and other metals in porewater due to the presence of wetland plant roots. Although, porewater concentrations of metals and S of mine impacted wetlands have been found to increase throughout the active growing phase in the presence of Phragmites australis (Batty et al. 2006). The results of the present study have also confirmed that the presence of the tested species resulted in increased porewater concentrations of metals and S throughout the active growing phase. Porewater concentrations of Cu, Mg, Na, S and Zn increased significantly over the trial period, due to radial oxygen loss. Increased concentrations in porewater is known to occur due to evapotranspiration during warmer summer months; release of oxygen into the anoxic substrate can oxidize previously formed metal sulfides, releasing metals into porewater (Batty et al. 2006). Of particular importance for the present study is that increasing porewater concentrations impacted photosynthetic performance.

However, the significantly greater percentages of N found in porewater and leaf material appears to have been a key contributor to the ability of each species to cope with stressors to photosynthetic performance due to accumulation of excess concentrations of elements. The leachate, porewater and subsequent plant material contained significantly greater concentrations of N in treatment conditions compared with the control conditions. Mine waste rock–dumps have been found to not only be a source of metalliferous leachate, but also a source of nitrogen from residues of ammonium nitrate based explosives used in blasting the rock. The waste–rock can have surficial blasting residues, which can leach nitrogen into

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surface water. The nitrogen is expected to be in the form of nitrate, although both ammonia and nitrate forms are possible. Nitrite and ammonia are more toxic to aquatic life than nitrate, but react quickly with oxygen to form nitrate. These effects are expected to be short-lived and will continue only until the blasting residue is leached from the rock surface (Fitch and Thompson 2000); the size and quantity of waste rock will also determine the release period of N from the dump.

Several key outcomes have been identified throughout this study. The natural colonisation and role of T. domingensis in reducing concentrations of Cu, Mn, and

Zn in leachate confirms that this species is highly desirable for phytoremediation of saline mine–leachate in particular. The hydroponic assessment of B. articulata, C. appressa, P. australis, and S. validus in synthetic simulated leachate indicated each species showed promise for phytoremediation of saline mine–leachate. Although, once the performance of these species was assessed in natural conditions the impact of significantly impacted photosynthetic assimilation. It has been established in other studies and in the present study that the presence of wetland plants and subsequent oxidization of the rhizosphere can increase porewater concentrations of metals and sulphate. This study has however determined that an increase of porewater concentrations can significantly impact photosynthetic performance

(Figure 13). Analysis of porewater concentrations of Cu, Mg, Na, and S revealed the greatest concentrations were reached at the conclusion of the trial period. It is however apparent that the presence of additional N in the leachate enabled the plants to overcome the toxicity of Cu, Mg, Na, and S; although as the porewater

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concentrations continued to increase photosynthetic assimilation was significantly impacted towards the conclusion of the trial period.

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Reference list

Adams AA, Raman A, Hodgkins DS, Nicol HI. (2013) Accumulation of heavy metals by naturally colonising Typha domingensis (Poales: Typhaceae) in waste-rock dump leachate storage ponds in a gold-copper mine in the Central Tablelands of New South Wales, Australia. International Journal of Mining Reclamation, and Environment 27: 294–307. Adhikari T, Kumar A, Singh MV, Subba Rao A. (2010) Phytoaccumulation of lead by selected wetland species. Communications in Soil Science and Plant Analysis 41: 2623–2632. Adriano DC, Wenzel WW, Vangronsveld J, Bolan NS. (2004) Role of assisted natural remediation in environmental cleanup. Geoderma 122: 121–142. Aksoy A, Demirezen D, Duman F. (2005) Bioaccumulation, detection and analyses of heavy metal pollution in sultan marsh and its environment. Water Air Soil Pollution 164: 241–255. Almeida CMR, Dias AC, Mucha AP, Bordalo AA, Vasconcelos MTSD. (2009) Influence of surfactants on the Cu phytoremediation potential of a salt marsh plant 75: 35–140. Alloway BJ. (1990) Soil processes and the behavior of metals. In Heavy metals in soils. (ed. BJ Alloway). Blackie Academic & Professional. Glasgow. p. 7– 28. An J, Kim JY, Kim KW, Park JY, Lee JS, Jang M. (2011) Natural attenuation of arsenic in the wetland system around abandoned mining area. Environmental Geochemical Health 33: 71–80. Antonovics J, Bradshaw AD, Turner RG. (1971) Heavy metal tolerance in plants. Advances in Ecological Research 7: 1–85. Arfstrom C, Macfarlane AW, Jones RD. (2000) Distributions of mercury and phosphorous in Everglades soils from water conservation area 3A, Florida, U.S.A. Water, Air, Soil Pollution 121: 133–159. Armstrong W, Drew MC. (2002) Root growth and metabolism under oxygen deficiency. In Plant Roots: The Hidden Half (3rd ed). (eds Waisel Y, Eshel A, and Kafkafi U). Marcel Dekker. New York. p. 729–761. Armstrong J, Armstrong W, Beckett PM. (1992) Phragmites australis: Venturi- and humidity-induced pressure flows enhance rhizome aeration and rhizosphere oxidation. New Phytologist 120: 197–207. Armstrong W. (1975) Waterlogged soils. In Environment and Plant Ecology. (ed. JR Etherington). Wiley. London. p. 181-218. Armstrong W. (1978) Root aeration in wetland conditions. In Plant Life in Anaerobic Conditions (eds. Hook DD. and Crawford RMM). Ann Arbor Science. Ann Arbor. MI. p. 269–297. Armstrong J, Armstrong W. (1988) Phragmites australis – a preliminary study of soil oxidizing sites and internal gas transport pathways. New Phytologist 108: 373–382. Armstrong W. (1982) Waterlogged soils. In Environment and Plant Ecology. (ed. JR Etherington). John Wiley and Sons, Chichester, UK. p. 290–330.

161

Armstrong J, Armstrong W. (1990) Light-enhanced convective throughflow increases oxygenation in rhizomes and rhizosphere of Phragmites australis (Cav.) Trin. Ex Steud. New Phytologist 114: 128–128. Armstrong W, Justin SHF, Beckett PM, Lythe S. (1991) Root adaptation to soil waterlogging. Aquatic Botany 39: 57–73. Arthur EL, Rice PJ, Anderson TA, Baladi SM, Henderson KLD, Coats JR. (2005) Phytoremediation – An Overview. CRC Critical Reviews in Plant Science 24: 109–122. Ashraf MA, Maah MJ, Yusoff I. (2011) Heavy metal accumulation in plants growing in ex tin mining catchment. International Journal of Environmental Science Technology 8: 401–416. Athay D, Nairn R, Strevett K. (2003) Mine drainage treatment wetland substrate analysis. In Proceedings of the 20th Meeting of the American Society for Surface Mining and Reclamation (ed Barnhisel RI). American Society for Mining and Reclamation, Billings. Montana. p. 2–16. Atwell BJ, Kriedemann PE, Turnbull CGN (eds). (1999) Plants in Action: Adaption in Nature, Performance in Cultivation. Macmillian Education, South Yarra, Victoria, Australia. Bacha RE, Hossner LR. (1975) Characteristics of coatings formed on rice roots as affected by iron and manganese additions. Soil Science Society of America Journal 41: 931–935. Baker AJM. (1981) Accumulators and excluders-strategies in the response plants to heavy metals. Journal of Plant Nutrition 3: 643–654. Baker AJM. (1987) Metal tolerance. New Phytologist 106: 93–111. Baker AJM, Brooks RR. (1989) Terrestrial higher plants which hyperaccumulate metallic elements. A review of their distribution, ecology and phytochemistry. Biorecovery 1: 81–26. Baker AJM, McGrath SP, Reeves RD, Smith JAC. (2000). Metal hyperaccumulator plants: a review of the ecology and physiology of a biological resource for phytoremediation of metal-polluted soils. In Phytoremediation of Contaminated Soil and Water. (eds. Terry N, and Banuelos G). Lewis. Boca Raton. p. 85–107. Balasoiu CF, Zagury GJ, Deschênes L. (2001) Partitioning and speciation of chromium, copper, and arsenic in CCA-contaminated soils: Influence of soil composition. Science of the Total Environment. 280: 239–255. Baldantoni D, Alfani A, Tommasi PD, Bartoli G, Santo AVD. (2004) Assessment of macro and microelement accumulation capability of two aquatic plants. 130: 149–156. Bandou E, Lebailly F, Muller F Dulormne M, Toribio A, Chabrol J, Courtecuisse R, Plenchette C, Prin Y, Duponnois R, Thiao M, Sylla S, Dreyfus B, Bâ AM.(2006) The ectomycorrhizal fungus Scleroderma bermudense alleviates salt stress in seagrape (Coccoloba uvifera L.) seedlings. Mycorrhiza 16: 559–565. Bastian RK, Hammer DA. (1993) The use of constructed wetland for wastewater treatment and recycling. In Constructed Wetlands for Water Quality Improvement. (ed. Moshiri AG). Lewis. Chelsea. Michigan. p. 59–68.

162

Batty LC, AJM, Wheeler BD, Curtis CD. (2000) The effect of pH and plaque on the uptake of Cu and Mn in Phragmites australis (Cav.) Trin ex. Steudel. Annals of Botany 86: 647–653. Batty LC, Baker AJM, Wheeler BD. (2002) Aluminium phosphate uptake by Phragmites australis: The role of Fe, Mn and Al root plaques. Annals of Botany 89: 443–449. Batty LC. (2003) Wetland plants – more than just a pretty face? Land Contamination Reclamation 11: 173–180. Batty LC, Younger PL. (2004) Growth of Phragmites australis (Cav.) Trin Ex. Steudel in mine water treatment wetlands: effects of metal and nutrient uptake. Environmental Pollution 132: 85–93. Batty LC, Baker AJM, Wheeler BD. (2006) The effect of vegetation on porewater composition in a natural wetland receiving acid mine drainage. Wetlands 26: 40–48. Benner SG, Smart EW, Moore JN. (1995) Metal behavior during surface- groundwater interaction, Silver Bow Creek, Montana. Environmental Science Technology 29: 989–1000. Beining BA, Otte ML. (1996) Retention of metals originating from an abandoned lead―zinc mine by a wetland at Glendalough, Co. Wicklow. Biology Environmental 96: 117–126. Bizily SP, Kim T, Kandasamy MK, Meagher RB. (2003) Subcellular targeting of methylmercury lyase enhances its specific activity for organic mercury detoxification in plants. Plant Physiology 131: 463–471. Boers AM, Zedler JB. (2008) Stabilized water levels and Typha invasiveness, Wetlands. 28: 676–685. Bonanno G, Giudice R. (2010) Heavy metal bioaccumulation by the organs of Phragmites australis (Common Reed) and their potential use as contamination indicators. Ecological Indicators 10: 639–645. Bonanno G. (2011) Trace element accumulation and distribution in the organs of Phragmites australis (common reed) and biomonitoring applications. Ecotoxicology and Environmental Safety 74: 1057–1064. Boye M, Van Den Berg CMG. (2000) Iron availability and the release of iron- complexing ligands by emiliania huxleyi. Marine Chemistry 70: 277–287. Bower C, Medd R. (1995) Flora report for Newcrest Mining on the Cadia Project, Technical Report, Newcrest Mining Limited, Orange. Bradford MM. (1976) A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Annals of Biochemistry 72: 248–254. Brady KU, Kruckeberg AR, Bradshaw Jr HD. (2005) Evolutionary ecology of plant adaptation to serpentine soils. Annual Review of Ecology, Evolution, and Systematics 36: 243–266. Brams E, Anthony W, Weatherspoon L. (1989) Biological monitoring of an agricultural food chain: Soil cadmium and lead in ruminant tissues. Journal of Environmental Quality 18: 317-323.

163

Bratieres K, Fletcher TD, Deletic A, Zinger Y. (2008) Nutrient and sediment removal by stormwater biofilters: A large-scale design optimisation study. Water Research 42: 3930–3940. Briggs BG, Johnson LAS. (1968) The status and relationships of the Australian species of Typha. Contributions from the New South Wales National Herbarium 4: 57–69. Bragato C, Brix H, Malagoli M. (2006) Accumulation of nutrients and heavy metals in Phragmites australis (Cav.) Trin. Ex Steudal and Bolboschoenus maritumus (L.) Palla in a constructed wetland of the Venice lagoon watershed. Environmental Pollution 144: 967–975. Brix H Sorrell BK, Orr PT. (1992) Internal pressurization and convective gas flow in some emergent freshwater macrophytes. Limnology Oceanography 37: 1420–1433. Brooks RR. (1991) Progressive development of wetland habitats at the RGC wetlands centre capel. Western Australia. Technical report No.17. Brooks RR, Chambers MF, Nicks LJ, Robinson BH. (1998) Phytomining. Trends in Plant Science 3: 359-362. Cabrera-Bosquet L, Albrizio R, Luis Araus J, Nogués S. (2009) Photosynthetic capacity of field-grown durum wheat under different N availabilities: a comparative study from leaf to canopy. Environmental and Experimental Botany 67: 145–152. Caçador I, Vale C, Catarino F. (2000) Seasonal variation of Zn, Pb, Cu and Cd concentrations in the root–sediment system of Spartina maritima and Halimione portulacoides from Tagus estuary salt marshes. Marine Environmental Research 49: 279–290. Caffrey JM, Kemp WM. (1991) Seasonal and spatial patterns of oxygen production, respiration and root-rhizome release in Potamogeton perfoliatus L. and Zostera marina. Aquatic Botany 40: 109–128. Cakmak I, Marshner H. (1993) Plant nutrition from genetic engineering field practice. In Plant nutritional status on superoxide radical and hydrogen peroxide scavenging enzymes in bean leaves. (Ed. NJ Barriow). Kluwer: Netherlands. P. 133–137. Caldelas C, Dong S, Araus JL, Weiss DJ. (2011) Zinc isotopic fractionation in Phragmites australis in response to toxic levels of zinc. Journal of Experimental Botany 62: 2169–78. Cannon HL, Connally GG, Epstein JB, Parker JG, Thornton I, Wixson G. (1978) Rocks: geological sources of most trace elements. In: report to the workshop at south scas plantation Captiva Island, FL, US. Geochemical Environments 3:17–31. Cao B, Shi L, Brown RN, Xiong Y, Fredrickson JK, Romine MF, Marshall MJ, Lipton MS, Beyenal H. (2011) Extracellular polymeric substances from shewanella sp. HRCR-1 biofilms: Characterization by infrared spectroscopy and proteomics. Environmental Microbiology 13: 1018–1031. Castelo-Grande T, Barbosa D. (2003) Soil decontamination by supercritical extraction. Journal of Environment, Agriculture and Food Chemistry 2: 331–336.

164

Chaffei C, Pageau K, Suzuki A, Gouia H, Ghorbel MH, Masclaux-Daubresse C. (2004) Cadmium toxicity induced changes in nitrogen management in Lycopersicon esculentum leading to a metabolic safeguard through an amino acid storage strategy. Plant and Cell Physiology 45: 1681–1693 Chaney RL, Malik M, Li YM, Brown SL, Brewer EP, Angle JS, Baker AJM. (1997) Phytoremediation of soil metals. Current Opinion in Biotechnology 8: 279-284. Chen CC, Dixon JB, Turner FT. (1980a) Iron coatings on roots: mineralogy and quantity influencing factors. Soil Science society of America Journal 44: 635–639. Christensen PB, Revsbech NP, Sand-Jensen K. (1994) Microsensor analysis of oxygen in the rhizosphere of the aquatic macrophyte Littorella uniflora (L) Ascherson. Plant Physiology 105: 847–852. Clemente R, Dickinson NM, Lepp NW. (2008) Mobility of metals and metalloids in a multi-element contaminated soil 20 years after cessation of the pollution source activity. Environmental Pollution 155: 254–261. Cobbett CS. (2000) Phytochelatins and their roles in heavy metal detoxification. Plant Physiology 123: 825–832. Connell EL, Colmer TD, Walker DI. (1999) Radial oxygen loss from intact roots of Halophila ovalis as a function of distance behind the root tip and shoot illumination. Aquatic Botany 63: 219–228. Connick VJ, Raman A, Hodgkins DS, Simmons AT, Nicol HI. (2010) Contents of Al, Cu, Fe, and Mo in phalaris aquatica and trifolium subterraneum grown on an unamended and amended gold-mine tailings-storage facility in central-western New South Wales, Australia. International Journal of Mining, Reclamation and Environment 24: 195–205. Coplen TB. (1996). New guidelines for reporting stable hydrogen, carbon, and oxygen isotope-ratio data. Geochimica et Cosmochimica Acta 60: 3359–3360. Corami A, Mignardi S, Ferrini V. (2007) Copper and zinc decontamination from single- and binary-metal solutions using hydroxyapatite. Journal of Hazardous Materials 146: 164–170. Cunningham SD, Ow DW. (1996) Promises and prospects of phytoremediation. Plant Physiology 110: 715-719. Davis L, Castro-Diaz s, Azhang Q, Erickson L. (2002) Benefits of vegetation for soils with organic contaminants. Critical reviews in Plant Science 21: 457– 491. Davies LC, Dana J, Selk K, Smith K, Goplen B, Erikson LE. (1998) Movement of chlorinated solvents and other volatile organics through plants monitored by fourier transform infrarred (FT-IR) Spectrometry. Journal of Hazardous Substances Research 1: 1–26. Davies FT, Puryear JD, Newton RJ, Egilla JN, Grossi JAS. (2002) Mycorrhizal fungi increase chromium uptake by sunflower plants: influence on tissue mineral concentration, growth and gas exchange. Journal of Plant Nutrition 25: 2389–2407.

165

Davy DR. (1975) Rum Jungle Environmental Studies. Australian Atomic Energy Commision, Lucas Heights, NSW. Davydova SL. (1999) Dangerous lead pollution in Russia. Russian Journal of Organic Chemistry 35: 160-163. Demmig-Adams B, Adams WW. (1992) III Photoprotection and other responses of plants to high light stress. Annual Reviews in Plant Physiology and Plant Molecular Biology 43: 599–626. Demirevska-Kepova K, Simova-Stoilova L, Stoyanova Z, Hölzer R, Feller U. (2004) Biochemical changes in Barley plants after excessive supply of copper and manganese. Environmental Experimental Botany 52: 253–266. Deng CH, Zhang G, Pan X. (2011) Photosynthetic responses in Reed (Phragmites australis (CAV.) TRIN. Ex Steud.) seedlings induced by different salinity- alkalinity and nitrogen levels. Journal of Agricultural Science Technology 13: 687–699. Deng CN, Zhang GX, Pan XL, Zhao KY. (2010) Chlorophyll fluorescence and gas exchange responses of Maize seedlings to saline-alkaline stress. Bulgarian Journal of Agricultural Science 16: 49–58. Deng H, Ye ZH, Wong MH. (2004) Accumulation of lead, zinc, copper and cadmium by twelve wetland plant species thriving in metal contaminated sites in China. Environmental Pollution 132: 29–40. Denny P. (1972) Sites of nutrient absorption in aquatic macrophytes. Journal of Ecology 60: 819–29. Dermont G, Bergeron M, Mercier G, Richer-Laflèche M. (2008) Metal- contaminated soils: remediation practices and treatment technologies. Periodical of Hazardous, Toxic and Radioactive Waste Management 12, 188–209. Dickinson NM, Baker AJM, Doronila A, Laidlaw S, Reeves RD. (2009) Phytoremediation of inorganics: realism and synergies. International Journal of Phytoremediation 11: 97–114. Dhir B, Sharmila P, Pardha Saradhi P, Sharma S, Kumar R, Mehta D. (2011) Heavy metal induced physiological alterations in Salvinia natans. Ecotoxicology and Environmental Safety 74: 1678–1684. Doyle M, Otte M. (1997) Organism-induced accumulation of iron, zinc, and arsenic in wetland soils. Environmental Pollution 96: 1–11 1997. Drever JL. (1988) The geochemistry of natural waters. Prentice-Hall: Englewood Cliffs. Dudeney AWL, Chanb BKC, Bouzalakos S, Huisman JL. (2012) Management of waste and wastewater from mineral industry processes, especially leaching of sulphide resources: state of the art. International Journal of Mining, Reclamation and Environment 27: 1–36. Dunbabin JS, Bowmer KH. (1992) Potential use of constructed wetlands for treatment of industrial wastewaters containing metals. Science of the Total Environ 111: 151–168. Dudka S, Adriano DC. (1997) Environmental impacts of metal ore mining and processing: a review. Journal of Environmental Quality 26: 590–602.

166

Dunbabin JS, Bowmer KH. (1992) Potential use of constructed wetlands for treatment of industrial wastewaters containing metals. Science of the Total Environment 111: 151–168. Eger P. (1994) Wetland treatment for trace metal removal from mine drainage: the importance of aerobic and anaerobic processes. Water and Science Technology 29: 249–256. Egiebor NO, Oni B. (2007) Acid rock drainage formation and treatment: a review. Asia-Pacific Journal of Chemical Engineering 2: 47–62. Elmayan T, Tepfer M. (1994) Synthesis of a bifunctional metallothionein/beta- glucuronidase fusion protein in transgenic tobacco plants as a means of reducing leaf cadmium levels. Plant Journal 6: 433–440. Ensley BD. (2000) Rationale for use of phytoremediation. In Phytoremediation of Toxic Metals: Using Plants to Clean up the Environment. (eds. Raskin I, Ensley BD). John Wiley and Sons, Inc, New York. p. 1–12. Ernst WHO. (1996) Bioavailability of heavy metals and decontamination of soils by plants. Applied Geochemistry 11: 163–167. Ernst WHO. (2005) Phytoextraction of mine wastes – options and impossibilities. Chemie der Erde Geochemistry 65: 29–42. Evangelou VP, Zhang YL. (1995) A review: pyrite oxidation mechanisms and acid mine drainage prevention. Critical Reviews of Environmental Science Technology 25: 141–199. Evans JR. (1989) Photosynthesis and nitrogen relationships in leaves of C3 plants. Oecologia 78: 9–19. Evans KM, Gatehouse JA, Lindsay WP, Shi J, Tommey AM, Robinson NJ. (1992) Expression of the Pea metallothionein-like gene PsMTA in Escherichia coli and Arabidopsis thaliana and analysis of trace metal ion accumulation: implications for PsMTA function. Plant Molecular Biology 20: 1019–1028. Feller U, Anders I, Demirevska K. (2008) Degradation of rubisco and other chloroplast proteins under abiotic stress. General Applied Plant Physiology 34: 5–18. Field C, Mooney HA. (1986) The photosynthesis-nitrogen relationship in wild plants. In On the Economy of Form and Function. (ed. Givinsh TJ.) Cambridge University Press. Cambridge. UK. p. 25–55. Fitch M, Thompson M. (2000) The environmental effects of rock drains. In Environmental Issues and Management of Waste in Energy and Mineral Production. (eds. Singhal RK, Mehrotra AK). A.A. Balkema, Rotterdam, Netherlands. Flanagan LB, Jefferies RL. (1989) Effect of increased salinity on CO2 assimilation 13 O2 evolution and the d C values of leaves of Plantago maritima L., developed under low and high NaCI levels. Planta 178: 377–384. Fulekar MH, Singh A, Bhaduri AM. (2009) Genetic engineering strategies for enhancing phytoremediation of heavy metals. African Journal of Biotechnology 8: 529–535. Fuller CC, Harvey JW. (2000) Reactive uptake of trace metals in the hyporheic zone of a mining-contaminated stream, Pinal creek, Arizona. Environmental Science Technology 34: 1150–1155.

167

Furrer G, Phillips BL, Ulrich KU, Pöthig R, Casey W. (2002) The origin of aluminum flocs in polluted streams. Science 297: 2245–2247. Galletti A, Verlicchi P, Ranieri E. (2010) Removal and accumulation of Cu, Ni and Zn in horizontal subsurface flow constructed wetlands: contribution of vegetation and filling medium. Science of the Total Environment 408: 5097–5105. Galmés J, Aranjuelo I, Medrano H, Flexas J. (2013) Variation in Rubisco content activity under variable climatic factors. Photosynthesis research (DOI 10.1007/s11120-013-9861-y). Gambrell RP. (1994) Trace and toxic metals in wetlands – a review. Journal of Environmental Quality 23: 883–891. Genty B, Briantais JM, Baker NR. (1989) The relationship between the quantum yield of photosynthetic electron transport and quenching of chlorophyll fluorescence. Biochimica et Biophysica Acta 990: 87–92. Ghosh M, Singh SP. (2005) A review on phytoremediation of heavy metals and utilization of its byproducts. Applied Ecology Environmental Research 3: 1–18. Gilbert N. (2010) Mountain mining damages streams. Nature 466: 806. Glass DJ. (2000) Economic potential of phytoremediation. In Phytoremediation of Toxic Metals: Using Plants to Clean up the Environment (eds. Raskin I, Ensley BD). John Wiley & Sons, Inc, New York. p. 15–31. Guilizzoni P. (1991) The role of heavy metals and toxic materials in the physiological ecology of submersed macrophytes. Aquatic Botany 41: 87– 109. Gräfe M, Naidu R. (2008) Remediation of metal contaminated sites: an overview. In Biophysico-Chemical Processes of Heavy Metal and Metalloids in Soil Environments. (eds. Violante A, Huang PM). John Wiley & Sons, Inc, Hoboken, NJ, USA. p. 565–605. Greger M. (2004) Metal availability, uptake, transport and accumulation in plants. In heavy metal stress in plants: from molecules to ecosystems. (2nd edn). (eds. Prasad MNV, Hagermeyer J). Springer Verlag, Heidelberg, Germany. p. 1–27. Guilizzoni P. (1991) The role of heavy metals and toxic materials in the physiological ecology of submersed macrophytes. Aquatic Botany 41: 87– 109. González A, Lynch JP. (1997) Effects of manganese toxicity on leaf CO2 assimilation of contrasting common bean genotypes. Physiology of Plants 101: 872–880. Halford KJ. (1999) A design criteria for a management system for the control and treatment of acid-rock drainage from mining operations in the Northern Territory. Mineral Resources Engineering 8: 419–452. Hall AH. (2002) Chronic arsenic poisoning. Toxicology Letters 128: 69–72. Hansch R, Mendel RR. (2009) Physiological functions of mineral micronutrients (Cu, Zn, Mn, Fe, Mo, B, Ci). Current Opinions in Plant Biology 12: 259– 266.

168

Hares RJ, Ward NI. (2004) Sediment accumulation in newly constructed vegetative treatment facilities along a new major road. Science of the Total Environment 334-335: 473-479. Hasegawa PM. (2013) Sodium (Na+) homeostasis and salt tolerance of plants. Environmental and Experimental Botany 92: 19–31. Hawkesford M J & De Kok L J 2006 Managing sulfur metabolism in plants. Plant Cell and Environment 29: 382―395. Hegazy AK, Abdel-Ghani, NT, El-Chaghaby GA. (2011) Phytoremediation of industrial wastewater potentiality by Typha domingensis. International Journal of Environmental Science Technology 8: 639–648. Heisey RM, Damman AWH. (1982) Copper and lead uptake by aquatic macrophytes in eastern Connecticut, U.S.A. Aquatic Botany 14: 213–229. Hernandez-Allica J, Becerril JM, Zarate O, Garbisu C. (2006). Assessment of the efficiency of a metal phytoextraction process with biological indicators of soil health. Plant Soil 281: 147–158. Hitch M, Ballantyne SM, Hindle SR. (2010) Revaluing mine waste rock for carbon capture and storage. International Journal of Mining, Reclamation and Environment 24: 64–79. Hong S, Candelone J-P, Patterson CC, Boutron CF. (1996) History of ancient copper smelting pollution during roman and medieval times recorded in Greenland ice. Science 272: 246–249. Horne AJ. (2000) Phytoremediation by constructed wetlands. In: Phytoremediation of contaminated soil and water,(eds. Terry N, Bañuelos G. Boca Raton: Lewis. pp. 13-40. Horner R. (1995) Use of constructed wetlands for stormwater quality enhancement. In Professional engineering practice liaison program. University of Washington. Seattle. Houtz R, Nable RO, Cheniae GM. (1988) Evidence for effects on the in vivo activity of ribulose-biphosphate carboxylase/oxygenase during development of Mn toxicity in tobacco. Plant Physiology 86: 1143–1149. Hupfer M, Dollan A. (2003) Immobilisation of phosphorus by iron-coated rots of submerged macrophytes. Hydrobiologia 506–509 (1-3) 635–640. Li H, Ye ZH, Wei ZJ, Wong MH. (2011) Root porosity and radial oxygen loss related to arsenic tolerance and uptake in wetland plants. Environmental Pollution 159: 30–37. Mora ML, Rosas A, Ribera A, Rengel Z. (2009) Differential tolerance to Mn toxicity in perennial ryegrass genotypes: involvement of antioxidative enzymes and root exudation of carbohydrates 320: 79–89. Jackson LJ. (1998) Paradigms of metal accumulation in rooted aquatic vascular plants. Science of the Total Environment 219: 223-231. Jacob DL, Otte ML. (2003) Conflicting processes in the wetland plant rhizosphere: metal retention or mobilization? Water Air Soil Pollution 3: 91–104. Jacob DL and Otte ML. (2004) Long-term effects of submergence and wetland vegetation on metals in a 90-year old abandoned Pb―Zn mine tailings pond. Environmental Pollution 130: 337–345.

169

Jamal A, Ayub N, Usman M, Khan AG. (2002) Arbuscular mycorrhizal fungi enhance Zinc and Nickel uptake from contaminated soil by soybean and Lentil. International Journal of Phytoremediation 4: 205–221. Jensen ML, Bateman AM. (1979) Economic Mineral Deposits. John Wiley and Sons, New York. Jenssen PD, Maehlum T, Krogstad T. (1993) Potential use of constructed wetlands for wastewater treatment in northern environments. Water Science Technology 28: 149–157. Johnson DB, Hallberg KB. (2005) Acid mine drainage remediation options: a review. Science of the Total Environment 338: 3–14. Joner EJ, Leyval C. (1997) Uptake of 109cd by roots and hyphae of a Glomus mosseae/Trifolium subterraneum mycorrhiza from soil amended with high and low concentrations of cadmium. New Phytologist 135: 353–360. Johnston CA. (1993) Mechanisms of wetland water quality interaction. In Constructed wetland for water quality improvement. (ed. GA Moshiri). Lweis Publishers. Boca Raton. p. 293–300. Johnston J, Murray G, Ritchie I (1997) Managing sulphidic mine wastes and acid drainage. Australia. Justin SHFW, Armstrong W. (1987) The anatomical characteristics of roots and plant response to soil flooding. New Phytologist 105: 465-495. Juwarkar AA, Singh SK, Mudhoo A. (2010) A comprehensive overview of elements in bioremediation. Reviews of Environmental Science and Biotechnology 9: 215–288. Kayser A, Wenger K, Keller A, Attinger W, Felix HR, Gupta SK, Schulin R. (2000) Enhancement of phytoextraction of Zn, Cd, and Cu from calcareous soil, the use of NTA and sulfur amendments. Environment Science Technology 34: 1778–1783. Keller C, Hammer D, Kayser A, Richner W, Brodbeck M, Sennhauser M. (2003) Root development and heavy metal phytoextraction efficiency: comparison of different plant species in the field. Plant Soil 249: 67–81. Kelly J, Champagne P, Michel F. (2006) Mitigation of alkaline mine drainage in a natural wetland system. In Geo-Environment and Landscape Evolution II, (eds. Martin-Duque JF, Brebbia, CA, Emmanouloudis D). WIT Press. Southhampton. p. 115–24. Keraita B, Konradsen F, Drechsel P. (2006) Farm-based measures for reducing human and environmental health from chemical constituents of wastewater. In Wastewater Irrigation and Health: Assessing and Mitigating Risk in Low-income Countries. (eds. Drechsel P, Scott DA, Raschid-Sally L, Redwood M, and Bahri A). International Water Management Institute, International Development Research Centre, and Earthscan, London. p. 189–208. Khan S, Ahmad I, Shah MT, Rehman S, Khaliq A. (2009) Use of constructed wetland for the removal of heavy metals from industrial wastewater. Journal of Environmental Management 90: 3451–3457. Kidd P, Barceló J, Pilar Bernal M, Navari-Izzo F, Poschenrieder C, Shilev S, Clemente R, Monterroso C. (2009) Trace element behaviour at the root-soil

170

interface: Implications in phytoremediation. Environmental and Experimental Botany 67: 243–259. King N, Beard J, Gibbs A. (2003) Contamination of an upland stream by heavy metals from an old mine site. Australian Journal of Ecotoxicology 9: 61–68. Kiser T, Hansen J, Kennedy B. (2010) Impacts and pathways of mine contaminants to Bull Trout (Salvelinus confluentus) in an Idaho watershed. Archives of Environmental Contamination Toxicology 59: 301–311. Kitao M, Lei TT, Koike T. (1997) Comparison of photosynthetic responses to manganese toxicity of deciduous broad-leaved trees in northern Japan. Environmental Pollution 97: 113–118. Kleinmann RLP. (1985) Treatment of acid mine water by wetlands. In, Control of acid mine drainage. Bureau of Mines Inf. Circ. 9027. U. S. Dep. of the Interior, Washington DC. pp. 17–35. Knight B, Zhao FJ, McGrath SP, Shen ZG. (1997) Zinc and cadmium uptake by the hyperaccumulator Thlaspi caerulescens in contaminated soils and its effects on the concentration and chemical speciation of metals in soil solution. Plant Soil 197: 71–78. Köhl K.I. and Lösch R. (2004) Experimental charecterisation of metal tolerance, in Chapter 17 ‘Heavy metals stress in plants: from biomolecules to ecosystems. MNV. Prasad, Hagemeyer J. (eds.) 2nd edn. Springer-Verlag, New York, pp. 371–389. Kotrba P, Najmanova J, Macek T, Ruml T, Mackova M. (2009) Genetically modified plants in phytoremediation of heavy metal and metalloid soil and sediment pollution. Biotechnological Advances 27: 799–810. Kovacs M. (1978) The element accumulation in submerged aquatic plant species in lake Balaton. ACTA Botanica Hungarica 24: 273–283. Kovacs M, Nyary I, Toth L. (1984) The microelement content of some submerged and floating aquatic plants. ACTA Botanica Hungarica 30: 173–185. Lambeck RJ. (2009) Mine closure or mind closure – are mining companies meeting their whole of life cycle, triple bottom line obligations? In Fourth International Conference on Mine Closure. (eds. Fourie AB, Tibbett M). Australian Centre for Geomechanics, The University of Western Australia, Perth. p. 13–19. Lasat MM. (2000) Phytoextraction of metals from contaminated soil: a review of plant/soil/metal interaction and assessment of pertinent agronomic issues. Journal of Hazardous Substances Research 2: 1–25. Lasat MM. (2002) Phytoextraction of toxic metals: a review of biological mechanisms. Journal of Environmental Quality 31: 109–120. LeDuc DL, Tarun DS, Montes-Bayon M, Meija M, Malit MF, Wu CP, Abdelsamie M, Chiang CY, Tagmount A, deSouza M, Neuhierl B, Böck A, Caruso J, Terry N. (2004) Overexpression of selenocysteine methyltransferase in Arabidopsis and Indian Mustard increases selenium tolerance and accumulation. Plant Physiology 135: 377–383. LeDuc D.L, AbdelSamie M, Mãntes-Bayon M, Wu CP, Reisinger SJ, Terry N. (2006) Overexpressing both ATP sulfurylase and selenocysteine

171

methyltransferase enhances selenium phytoremediation traits in Indian Mustard. Environmental Pollution 144: 70–76. Levitt J. (1980) Responses of Plants to Environmental Stresses. Academic Press, New York. Levine SN, Rudnick DT, Kelly JR, Morton RD, Buttel LA. (1990) Pollutant dynamics as influenced by seagrass beds: experiments with tributyltin in Thalassia microcosms. Marine Environment Research 30: 297–322. Lichaz W, Myers S. (1977) Uranium mining in Australia. In Ground for Concern – Australia’s Uranium and Human Survival (ed. Elliott M). Penquin. Melbourne. p. 25–63. Lidon FC, Barreiro MG, Ramalho JC. (2004) Manganese accumulation in rice: implications for photosynthetic functioning. Journal of Plant Physiology 144: 1235–1244. Li S, Mendelssohn IA, Chen H, Orem WH. (2009) Does sulphate enrichment promote the expansion of Typha domingensis (cattail) in the Florida Everglades? Freshwater Biology 54: 1909–1923. Lindsay WL, Norvell WA. (1978) Development of DTPA soil test for zinc, iron, manganese and copper. Soil Science Society of America Journal 42: 421– 428. Liu JR, Suh MC, Choi D. (2000) Phytoremediation of cadmium contamination: overexpression of metallotionein in transgenic Tobacco plants. Bundesgesundheitsblatt Gesundheitsforschung Gesundheitsschutz: 43: 126– 130. Liu WJ, Zhu YG, Smith FA, Smith SE. (2004) Do phosphorus nutrition and iron plaque alter arsenate (As) uptake by rice seedlings in hydroponic culture? New Phytologist 162: 481–488. Lombi E, Zhao FJ, Dunham SJ, McGrath SP. (2001) Phytoremediation of heavy metal-contaminated soils: natural hyperaccumulation versus chemically enhanced phytoextraction. Journal of Environmental Quality 30: 1919– 1926. MacFarlane GR, Burchett MD. (2002) Toxicity, growth and accumulation relationships of copper, lead and zinc in the grey mangrove avicennia marina (Forsk.) Vierh. Marine Environment Research 54: 65–84. Macnair MR, Baker AJM. (1994) Metal-tolerant plants: an evolutionary perspective. in Farago, M.E. (ed). plants and the chemical elements – biochemistry, uptake, tolerance and toxicity 67–85. VCH, Weinheim. Macnair MR, Smirnoff N. (1999) Use of zincon to study uptake and accumulation of zinc tolerant and hyperaccumulating plants. Communications in Soil Science and Plant Analysis 30: 1127–1136. Maddocks GA. (2009) Reactive In-Situ Covers for the Remediation of Mine Waste Rock Using BAUXSOL Lime and Biosolids. Southern Cross University, PhD thesis, Southern Cross University, Australia. Madhavan S, Treichel I, OLeary MH. (1991) Effects of relative humidity on carbon isotope fractionation in plants. Botanica Acta 104: 292–294. Marchand L, Mench M, Jacob DL, Otte ML. (2010) Metal and metalloid removal in constructed wetlands, with emphasis on the importance of plants and

172

standardized measurements: a review. Environmental Pollution 158: 3447– 3461. Mariotti A. (1983) Atmospheric nitrogen is a reliable standard for natural 15N abundance measurements. Nature 303: 685–687. Matagi SV, Swai D, Mugabe R. (1998) A review of heavy metal removal mechanisms in wetlands. African journal for tropical hyrdrobiology and fisheries 8: 23–35. Matsui Inoue T, Tsuchiya T. (2008) Interspecific differences in radial oxygen loss from the roots of three Typha species. Journal of Limnology 9: 207–211. Matthews DJ, Moran BM, Otte ML. (2005) Screening the wetland plant species Alisma plantago-aquatica, Carex rostrata and Phalaris arundinacea for innate tolerance to zinc and comparison with Eriophorum angustifolium and Festuca rubra merlin. Environmental Pollution 134: 343–351. Maine MA, Suñe N, Hadad H, Sánchez G, Bonetto C. (2009) Influence of vegetation on the removal of heavy metals and nutrients in a constructed wetland. Journal of Environmental Management 90: 355–363. Major D, Cox, E. (1992) Survey of Microbial Inoculants for Bioremediation and Identification of Information Requirements Suitable for the Feasibility Evaluation and Validation of Bioremediation. Hazardous contaminants branch of the Ontario ministry of the environment. queens printer for Ontario. USA. p. 216. Mayes WM, Batty LC, Younger PL, Jarvis AP, Kõiv M, Vohla C, Mander U. (2009) Wetland treatment at extremes of pH: a review. Science of the Total Environment. 407: 3944–3957. Maxwell K, Johnson GN. (2000) Chlorophyll fluorescence –a practical guide. Journal of Experimental Botany 51: 659–668. McCabe OM, Otte ML. (2000) The wetland grass Glyceria fluitans for revegetation of mine tailings. Wetlands 20: 548–559. McKee KL, Mendelssohn IA, Hester MW. (1988) Examination of pore water sulfide concentrations and redox potentials near the aerial roots of Rhizophora mangle and Avicennia germinans. American Journal of Botany 75: 1352–9. McGrath SP, Sidoli CMD, Baker AJM, Reeves RD. (1993) The potential for the use of metal-accumulating plants for the In Situ decontamination of metal- polluted soils. In Integrated Soil and Sediment Research: A Basis for Proper Protection. (eds. Eijsackers HJP, Hamers T). Kluwer Academic Publishers, Dordrecht, Netherlands. p. 673–676. McGrath SP, Lombi E, Gray CW, Caille N, Dunham SJ, Zhao FJ. (2006) Field evaluation of Cd and Zn phytoextraction potential by the hyperaccumulators thlaspi caerulescens and Arabidopsis halleri. Environmental Pollution 141: 115–125. Medina VF, McCutcheon SC. (1996) Phytoremediation: modeling removal of TNT and its breakdown products. Remediation 7: 31–45. Memon AR, Schröder P. (2009) Implications of metal accumulation mechanisms to phytoremediation. Environmental Science Pollution Research. 16: 162–175.

173

Mejáre M, Bülow L. (2001) Metal-binding proteins and peptides in bioremediation and phytoremediation of heavy metals. Trends in Biotechnology 19: 67–73. MESH Environmental Inc. (2009) Cadia East Project: Geochemical Characterization of Mine Rock and Tailings, Tech. Rep. M022–001, Mesh Environmental, Vancouver. Mehta P, Jajoo A, Mathur S, Bharti S. (2010). Chlorophyll a fluorescence study revealing effects of high salt stress on Photosystem II in wheat leaves. Plant Physiology and Biochemistry 48: 16–20. Miao SL, DeBusk WF. (1999) Effects of phosphorus enrichment on structure and function of sawgrass and cattail communities in the Everglades. In Phosphorus Biogeochemistry in Subtropical Ecosystems. (eds. K.R. Reddy, G.A. O’Connor, and C.L. Schelske). CRC Press, Boca Raton, Florida. p. 275–99. Migeon A, Richaud P, Guinet F, Blaudez D, Chalo M. (2012) Hydroponic screening of poplar for trace element tolerance and accumulation. International Journal of Phytoremediation 14: 350–361. Miller GE, Wile I, Hitchin GG. (1983) Patterns of accumulation of selected metals in members of the soft-water macrophyte flora of central Ontario lakes. Aquatic Botany 15: 53–64. Miller RM, Smith CR, Jastrow JD, Bever JD. (1999) Mycorrhizal status of the genus Carex (Cyperaceae). American Journal of Botany 86: 547–553. Misra S, Gedamu L. (1989) Heavy metal tolerant transgenic brassica napus L. and nicotiana tabacum L. plants. Theories in Applied Genetics 78: 161–168. Mitsch WJ, Gosselink JG (1993) Wetlands. Van Nostrand Reinhold: New York. Moreno-Jiménez E, Beesley L, Lepp NW, Dickinson NM, Hartley W, Clemente R. (2011) Field sampling of soil pore water to evaluate trace element mobility and associated environmental risk. Environmental Pollution 159: 3078– 3085. Mudd G. (2009) The sustainability of mining in Australia: key production trends and their environmental implications for the future, Monash University. Melbourne. Australia. Mudd GM, Patterson J. (2010) Continuing pollution from the rum jungle U-Cu project: a critical evaluation of environmental monitoring and rehabilitation. Environmental Pollution 158: 1252–1260. Mufarrege MM, Di Luca GA, Hadad HR, Maine MA. (2011) Adaptability of Typha domingensis to high pH and salinity. Ecotoxicology 20: 457–465. Mukhopadhyay S, Maiti SK. (2010) Phytoremediation of metal mine waste. Applied Ecology Environment Research 8: 207–222. Muller G. (1988) Chemical decontamination of dredged materials, combustion residues, soil and other materials contaminated with heavy metals. In The II international TNO/BMFT conference on contaminated soil. (Eds K Wolf, WJ Van, DE Brink and FJ Colon). Kluwer Dorrecht. p. 1439-1442.

174

Mulligan CN, Yong RN, Gibbs BF. (2001) Remediation technologies for metal- contaminated soils and groundwater: an evaluation. Engineering Geology 60: 193–207. Naftz DL, Rice JA. (1989) Geochemical processes controlling selenium in ground water after mining, Powder River Basin, Wyoming, U.S.A. Applied Geochemistry. 4: 565–576. Neelima P, Reddy KJ. (2004) Interaction of some heavy metals stress on growth and yield in solanum melongena L. Ecology, Environment, and Conservation 10: 153–159. Nagajyoti PC, Lee KD, Sreekanth SVM. (2010) Heavy metals, occurrence and toxicity for plants: a review. Environmental Chemistry Letters 8: 199-216. Neidorf R. (1996) Phytoremediation: phytotech’s hungry plants take up toxic metals. El digest, industrial and waste management, 4/18-21, 7–11. Neumann D, Zur Nieden U. (2001) silicon and heavy metal tolerance of higher plants. Phytochemistry 56: 685–692. Neumann PM, De Souza MP, Pickering IJ, Terry N. (2003) Rapid microalgal metabolism of selenate to volatile dimethylselenide. Plant Cell Environ. 26: 897–905. Naumann JC, Young DR, Anderson JE. (2008). Leaf chlorophyll fluorescence, reflectance, and physiological response to freshwater and saltwater flooding in the evergreen shrub, Myrica cerifera. Environmental and Experimental Botany 63: 402–409. Newschaffer CJ, Croen LA, Daniels J, Giarelli E, Grether JK, Levy SE, Mandell DS, Miller LA, Pinto-Martin J, Reaven J, Reynolds AM, Rice CE, Schendel D, Windham GC. (2007) The epidemiology of autism spectrum disorders. Annual Reviews of Public Health 28: 235–258. Noller BN, Woods PH, Ross BJ. (1994) Case studies of wetland filtration of mine waste water in constructed and naturally occurring systems in Northern Australia. Water Science and Technology 29: 257–265. Nordstrom DK, Alpers CN. (1999) Negative ph, efflorescent mineralogy, and consequences for environmental restoration at the iron mountain superfund site, California. Proceedings of the National Academy of Sciences U.S.A 96: 3455–3462. OKC (2010) South waste rock dump closure planning support studies - Newcrest Cadia Valley Operations. O’Leary MF. (1988) Carbon isotopes in photosynthesis. Bioscience 38: 328–335. Osanloo M, Rashidinejad F, Rezai B. (2008) Incorporating environmental issues into optimum cut-off grades modeling at porphyry copper deposits. Resources Policy 33: 222–229. O’Sullivan AD, Murray A, Otte ML. (2003) Constructed wetlands for treating processed mine water – an Irish case study. In Proceedings of the 3rd International Mining and Environment Meeting. (eds.Spiers G, Beckett P.M. and Conroy H). Canadian Land Reclamation Association. Sudbury. Ontario. Canada. p. 739–747.

175

O’S ullivan AD, Murray DA, Otte ML. (2004) Removal of sulfate, zinc, and lead from alkaline mine wastewater using pilot-scale surface-flow wetlands at Tara mines, Ireland. Mine Water Environ 23: 58–65. Otte ML. (2001) What is stress to a wetland plant? Environmental Experimental Botany 46: 195–202. Padmavathiamma PK, Li LY. (2007) Phytoremediation technology: hyperaccumulation metals in plants. Water Air Soil Pollution 184:105–126. Pagenkopf GK. (1978) Introduction to Natural Water Chemistry. Marcel Dekker, New York. Pavel LV, Gavrilescu M. (2008) Overview of ex-situ decontamination techniques for soil cleanup. Environmental Engineering and Management Journal 7: 815–836. Peuke AD, Rennenberg H. (2005) Phytoremediation. EMBO Reports 6: 497–501. Pietrini F, Iannelli M, Pasqualini S, Massacci A. (2003) Interaction of cadmium with glutathione and photosynthesis in developing leaves and chloroplasts of Phragmites australis (Cav.) Trin. ex Steudal. Plant Physiology 133: 829–837. Peer WA, Baxter IR, Richards EL, Murphy AS. (2005) Phytoremediation and hyperaccumulator plants. In: Klomp LWJ, Martinoia E, and Tamás MJ (Eds). Molecular biology of metal homeostasis and detoxification: from microbes to man. Topics in current genetics. New York, NY: Springer- Verlag. Peltier EF, Webb SM, Gaillard JF. (2003) Zinc and lead sequestration in an impacted wetland system. Advances in Environmental Research 8 103-112. Pilon-Smits E. (2005) Phytoremediation. Annual Reviews in Plant Biology 56: 15– 39. Plumlee GS, Smith KS, Montour MR, Ficklin WH, Mosier EL. (1999) Geologic controls on the composition of natural waters and mine waters draining diverse mineral- deposit types. In The Environmental Geochemistry of Mineral Deposits, Part B: Case Studies and Research Topics, Reviews in Economic Geology. (eds. Filipek LH, Plumlee GS). Society of Economic Geologists Inc, Colorado, USA. p. 373–432. Plumlee GS, Smith KS, Montour MR, Ficklin WH, Mosier EL. (2009) Geologic controls on the composition of natural waters and mine waters draining diverse mineral-deposit types. In The Environmental Geochemistry of Mineral Deposits, Part B: Case Studies and Research Topics, Reviews in Economic Geology. (eds. Filipek LH, Plumlee GS). Society of Economic Geologists, Park City, Utah. p. 373–432. Ponnamperuma FN. (1972) The chemistry of submerged soils. Advances in Agronomy 24: 29–96. Prasad MNV, Hagmeyer J. (1999) Heavy metal stress in plants. Springer: Berlin Prasad MNV, Freitas H, Fraenzle S, Wuenschmann S, Markert B. (2010) Knowledge explosion in phytotechnologies for environmental solutions. Environment Pollution 158: 18–23.

176

Qadir M, Oster, JD. (2004) Crop irrigation management strategies for saline-sodic soils and waters aimed at environmentally sustainable agriculture. Science of the Total Environment 323: 1–9. Rai PK. (2008) Heavy metal pollution in aquatic ecosystems and its phytoremediation using wetland plants: an ecosustainable approach. International Journal of Phytoremediation 10: 131–158. Rai PK. (2009) Heavy metal phytoremediation from aquatic ecosystems with special reference to macrophytes. Critical Reviews in Environmental Science Technology 39: 697–753. Rai PK. (2012) An eco-sustainable green approach for heavy metals management: two case studies of developing industrial region. Environmental Monitoring Assessments 184: 421–448. Rascio N, Navari-Izzo F. (2011) Heavy metal hyperaccumulating plants: how and why do they do it? and what makes them so interesting? Plant Science 180: 169–181. Reichamn SM. (2002) The Response of Plants to Metal Toxicity: A Review Focusing on Copper, Manganese and Zinc. the Australian minerals and energy environment foundation, Melbourne. Richards RJ, Applegate RJ, Ritchie AIM. (1996) The rum jungle rehabilitation project. In Environmental Management in the Australian Minerals and Energy Industries: Principles and Practices. (eds. Mulligan DR).. UNSW Press and Australian Minerals and Energy Environment Foundation (AMEEF). Sydney. p. 530–553. Robinson BH, Leblanc M, Petit D, Brooks RR, Kirkman JH, Gregg PEH. (1998) The potential of thlaspi caerulescens for phytoremediation of contaminated soils. Plant Soil 203: 47–56. Robinson D, Handley LL, Scrimgeour CM, Gordon DC, Forster BP, and Ellis RP. (2000) Using stable isotope natural abundances (δ15N and δ13C) to integrate the stress responses of wild barley (Hordeum spontaneum C. Koch.) genotypes. Journal of Experimental Botany 51: 41–50. Rugh CL, Senecoff JF, Meagher RB, Merkle SA. (1998) Development of transgenic yellow poplar for mercury phytoremediation. Nature Biotechnology 16: 925–928. Rulkens WH, Tichy R, Grotenhuis JTC. (1998) Remediation of polluted soil and sediment: Perspectives and failures. Water Science and Technology 37: 27– 35. Ruttens A, Colpaert JV, Mench M, Boisson J, Carleer R, Vangronsveld J. (2006) Phytostabilization of a metal contaminated sandy soil. II: Influence of compost and/or inorganic metal immobilizing soil amendments on metal leaching. Environmental Pollution 144: 533–539. Rybicka EH. (1996) Environmental impact of mining and smelting industries in Poland. In Environmental geochemistry and health with special reference to developing countries (eds. Appleton JD, Fuge R & McCall GJH). Geological Society. Special Publication. p. 183-193.

177

Ryan P, Hosking R. (1992) A Review of Wetland Filter Systems for Mining Operations. 17th Australian Mining Industry Council Environment Workshop. Yeppoon. Canberra. p.18. Salomons W. (1995) Environmental impact of metals derived from mining activities: Processes, predictions, prevention. Journal of Geochemical Exploration 52: 5–23. Salt DE, Blaylock M, Kumar NPBA, Dushenkov V, Ensley BD, Chet I, Raskin I. (1995) Phytoremediation: A novel strategy for the removal of toxic metals from the environment using plants. Bio/Technology 13: 468–474. Sarma H. (2011) Metal hyperaccumulation in plants: a review focusing on phytoremediation technology. Journal of Environmental Science Technology 4: 118–138. Sasaki K, Ogino T, Endo Y, Kurosawa K. (2003) Field study on heavy metal accumulation in a natural wetland receiving acid mine drainage. Material Transactions 44: 1877–1844. Schreiner RP, Mihara KL, McDaniel H, Bethlenfalvay GJ. (1997) Mycorrhizal fungi influence plant and soil functions and interactions. Plant Soil 188: 199–209. Shangguan ZP, Shao MA, Dyckmans J. (2000a) Effects of nitrogen nutrition and water deficit on net photosynthetic rate and chlorophyll fluorescence in winter wheat. Journal of Plant Physiology 156: 46–51. Sheoran AS, Sheoran V. (2006) Heavy metal removal mechanism of acid mine drainage in wetlands: A critical review. Minerals Engineering 19: 105-116. Shomar BH, Müller G, Yahya A. (2005) Seasonal variations of chemical composition of water and bottom sediments in the wetland of Wadi Gaza, Gaza Strip. Wetlands, Ecology, and Management 13: 419–431. Sholkovitz ER. (1978) The flocculation of dissolved Fe, Mn, Al, Cu, Ni, Co and Cd during esturine mixing. Earth Planet Science Letters 41: 77–86. Singer AC, Crowley DE, Thompson IP. (2003) Secondary plant metabolites in phytoremediation and bioaccumulation. Trends in Biotechnology 21: 123– 130. Singer OV, Stumm W. (1970) Acidic mine drainage; rate-determining step. Science 167: 1121–1123. Singh OV, Labana S, Pandey G, Budhiraja R, Jain RK. (2003) Phytoremediation: an overview of metallic ion decontamination from soil. Applied Microbiological Biotechnology 61: 405–412. Smith RAH, Bradshaw AD. (1979) The use of metal tolerant plant populations for the reclamation of metalliferous wastes. Journal of Applied Ecology 16: 595–612. Snowden RED, Wheeler BD. (1995) Chemical changes in selected wetland plant species with increasing Fe supply, with specific reference to root precipitates and Fe tolerance. New Phytologist 131: 503–520. Sobolewski A. (1999) A review of processes responsible for metal removal in wetlands treating contaminated mine drainage. International Journal of Phytoremediation 1: 19–51.

178

Stadtman ER, Oliver CN. (1991) Metal-catalyzed oxidation of proteins: physiological consequences. Journal of Biological Chemistry 266: 2005– 2008. Standards Association of Australia. AS 1289 c6.2-(1976) Determination of the particle size distribution of a soil: An analysis by sieving in combination with hydrometer analysis. St-cyr L, Campbell PGC. (1996) Metals (Fe, Mn, and Zn) in the root plaque of submerged aquatic plants collected in-situ: Relations with metal concentrations in the adjacent sediments and in the root tissue. Biogeochemistry 33: 45–76. St-Cyr, Crowder A.A. (1990) Manganese and copper in the root plaque of Phragmites australis (Cav. Trin. Ex Steudal. Soil Science 149: 191–198. St-Cyr L, Fortin D, Campbell PGC. (1993) Microscopic observations of the plaque of a submerged aquatic plant (Vallisneria Americana Michx). Aquatic Botany 46: 155–167. Stoltz E, Greger M. (2002) Accumulation properties of As, Cd, Cu, Pb and Zn by four wetland plant species growing on submerged mine tailings. Environmental Experimental Botany 47: 271–280. Stumm W, Morgan J. (1981) Aquatic chemistry. John Wiley & Sons: New York Strachotta C, Plewes H, McBrien M, Rey L. (2009) Mine closure. In Fourth International Conference on Mine Closure. (eds. Fourie AB and Tibbett M). Australian Centre for Geomechanics. The University of Western Australia. Perth. Australia. p. 285–97. Taylor GJ, Crowder AA, Rodden R. (1984) Formation and morphology of an iron plaque on the roots of Typha latifolia L. Grown in solution culture. American Journal of Botany 71: 666–675. Tsezos M, Remoudaki E, Angelatou V. (1997) Biosorption sites of selected metals using electron microscopy. Comparative Biochemistry and Physiology Part A Physiology 118: 481–487. Thompson EA, Luoma SN, Johansson CE, Cain DJ. (1984) Comparisons of sediments and organisms in identifying sources of biologically available trace metal contamination. Water Resources 18: 755–765. Tordoff GM, Baker AJM, Willis AJ. (2000) Current approaches to the revegetation and reclamation of metalliferous mine wastes. Chemosphere 41: 219-228. Towbin H, Staehelin T, Gordon J. (1979) Electrophoretic transfer of proteins from polyacrylamide gels to nitrocellulose sheets: procedures and some applications. Proceedings of the National Academy of Science, USA 76: 4350–4354. USEPA. (2009) Superfund green remediation strategy. U.S. Environmental Protection Agency. Vaillant N, Monnet F, Hitmi A, Sallanon H, Coudret A. 2005. Comparative study of responses in four Datura species to a zinc stress. Chemosphere 59:1005– 1013. Van Nevel L, Mertens J, Oorts K, Verheyen K. (2007) Phytoextraction of metals from soils: how far from practice? Environmental Pollution 150: 34–40.

179

Vangronsveld J, Herzig R, Weyens N, Boulet J, Adriaensen K, Ruttens A, Thewys T, Vassilev A, Meers E, Nehnevajova E, van der Lelie D, Mench M. (2009) Phytoremediation of contaminated soils and groundwater: lessons from the field. Environmental Science Pollution Research 16: 765–794. Vose PB, Randall PJ. (1962) Resistance to aluminum and manganese toxicities in plants related to variety and cation change capacity. Nature 196: 85–86. Vymazal J, Kröpfelová L. (2008) Wastewater Treatment in Constructed Wetlands with Horizontal Sub-Surface Flow. Springer, Dordrecht, The Netherlands. Vymazal J. (2011) Plants used in constructed wetlands with horizontal subsurface flow: A review. Hydrobiologica 674: 133–156. Walker CD, Sinclair R. (1992) Soil salinity is correlated with a decline in 13C discrimination in leaves of Atriplex species. Australian Journal of Ecology 17: 83–88. Wang Y, Inamori R, Kong H, Xu K, Inamori Y, Kondo T, Zhang J. (2008) Nitrous oxide emission from polyculture constructed wetlands: Effect of plant species. Environmental Pollution 152: 351–360. Wang W, Wang R, Yuan Y, Du1 N, Guo W. (2011) Effects of salt and water stress on plant biomass and photosynthetic characteristics of Tamarisk (Tamarix chinensis Lour.) seedlings. African Journal of Biotechnology 10: 17981- 17989. Watanabe ME. (1997) Phytoremediation on the brink of commercialization. Environmental Science Technology 31: 182–186. Wei S, Zhou Q, Srivastava M, Xiao H, Yanga C, Zhang Q. (2009) Kalimeris integrifolia Turez. Ex DC: an accumulator of Cd. Journal of Hazardous Materials 162: 1571–1573. Weider RK, Lang GE. (1982) Modification of acid-mine drainage in a freshwater wetland. In Symposium on Wetlands of the Unglaciated Appalachian Region. (ed. McDonald BR). West Virginia University. Morgantown. p. 43–53. Weibner A, Kappelmeyer U, Kuschk P, Kastner M. (2005) Influence of the redox condition dynamics on the removal efficiency of a laboratory-scale constructed wetland. Water Research 29: 248–256. Weis JS, Weis P. (2004) Metal uptake, transport and release by wetland plants: Implications for phytoremediation and restoration. Environment International 30: 685–700. Wielinga B, Lucy JK, Moore JN, Seastone OF, Gannon JE. (1999) Micro- biological and geochemical charecterisation of fluvially deposited sulphidic mine tailings. Applied Environmental Microbiology 65: 1548–1555. Wood, T.S. and Shelley, M.L. (1999) A dynamic model of bioavailability of metals in constructed wetland sediments. Ecological Engineering. 12: 231–252. Woodrow IE, Berry JA.(1988) Enzymatic regulation of photosynthetic CO2 fixation in C3 plants. Annual Review of Plant Physiology and Plant Molecular Biology 39: 33–594. Wong YS, Lam EKH, Tam NFY. (1994) Physiological effects of copper treatment and its uptake in Festuca rubra cv. merlin. Resources, Conservation and Recycling. 11: 311–19.

180

Woulds C, Ngwenya BT. (2004) Geochemical processes governing the performance of a constructed wetland treating acid mine drainage, Central Scotland. Applied Geochemistry 19: 1773–1783. Wright P, Mason CF. (1999) Spatial and seasonal variations in heavy metals in the sediments and biota of two adjacent estuaries, the Orwell and the Stour, in eastern England. Science of the Total Environment 226: 139–156. Wright DJ, Otte ML. (1999) Wetland plant effects on the biogeochemistry of metals beyond the rhizosphere. Biology and Environment, Proceedings of the Royal Irish Academy Volume 99B 1: 3-10. Wright, R.O. and Baccarelli, A. (2007) Metals and neurotoxicology. Journal of Plant Nutrition 137: 2809–2813. Wu, G., Kang, H., Zhang, X., Shao, H., Chu, L. and Ruan, C. (2010) A critical review on the bio-removal of hazardous heavy metals from contaminated soils: issues, progress, eco-environmental concerns and opportunities. Journal of Hazardous Materials 174: 1–8. Wu J, Wang L, Fang M, Yang J, Li S, Li Z. (2013) effects of vegetative-periodic- induced rhizosphere variation on the uptake and translocation of metals in Phragmites australis 9Cav.) Trin ex. Steudal growing in the Sun Island Wetland. Ecotoxicology 22: 608–618. Wu, S. (1994) Effect of manganese excess on the soybean plant cultivated under various growth conditions. Journal of Plant Nutrition 17: 991–1003. Yan H, Xiaotoa H, Li H. (2012) Leaf photosynthesis, chlorophyll fluorescence, ion content and free amino acids in Caragana korshinskii Kom exposed to NaCI stress. Acta Physiology Planta 34: 2285–2295 Yang X, Feng Y, He Z, Stoffella PJ. (2005) Molecular mechanisms of heavy metal hyperaccumulation and phytoremediation. Journal of Trace Elements in Medicine and Biology 18: 339–353. Ye ZH, Wong MH, Lan CY. (2004) Use of a wetland system for treating Pb/Zn mine effluent: a case study in southern China from 1984 to 2002. In Wetland Ecosystems in Asia: Function and Management. (ed. Wong MH). Elsevier, Amsterdam. p. 413–434. Ye ZH, Baker AJM, Wong MH, Willis AJ. (1998) Zinc, lead and cadmium accumulation and tolerance in Typha latifolia as affected by iron plaque on the root surface. Aquatic Botany 61: 55–67. Ye ZH, Baker AJM, Wong MH, Willis AJ. (2003) Copper tolerance, uptake and accumulation by phragmites australis. Chemosphere 50: 795–800. Yun S-I, Ro H-M. (2008) Stable C and N isotopes: a tool to interpret interacting environmental stresses on soil and plant. Journal of Applied Biology and Chemistry 56: 262–271. Ying R-R, Qiu R-L, Tang Y-T, Hu P-J, Qiu H, Chen H-R, Shi T-H, Morel J-L. (2010) Cadmium tolerance of carbon assimilation enzymes and chloroplast in Zn/Cd hyperaccumulator Picris divaricata. Journal of Plant Physiology 167: 81–87. Zayed A, Gowthaman S, Terry N. (1998) Phytoaccumulation of trace elements by wetland plants: i. duckweed. Journal of Environmental Quality 27: 715– 721.

181

Zhang YW, Tam NFY, Wong YS. (2004) Cloning and characterization of type 2 metallothionein-like gene from a wetland plant, typha latifolia. Plant Science 167: 869–877. Zhu YL, Pilon-Smits EAH, Jouanin L, Terry N. (1999) Overexpression of glutathione synthetase in indian mustard enhances cadmium accumulation and tolerance. Plant Physiology 119: 73–79. Zielinski RA, Otton JK, Johnson CA. (2001) Sources of salinity near a coal mine spoil pile, north-central Colorado. Journal of Environmental Quality 30: 1237–1248. Zinger Y, Blecken G-T, Fletcher TD, Viklander M, Deletić. (2013) Optimising nitrogen removal in existing stormwater biofilters: Benefits and tradeoffs of a retrofitted saturated zone. Ecological Engineering 51: 75–82.

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