An investigation into the use of Australian native plants and designer metal- binding particles for remediation of metal-contaminated landscapes Jaquelina Fátima Guterres Correia Master of Science
A thesis submitted for the degree of Doctor of Philosophy at The University of Queensland in 2018 School of Pharmacy Abstract
Contamination of soils with arsenic (As), lead (Pb), copper (Cu), manganese (Mn), zinc (Zn), cadmium (Cd) and aluminium (Al) as a result of mining and other industrial activities is a pervasive problem worldwide. High concentrations of these metal/loids and other contaminants in surface soil layers may prevent the growth of vegetation, leading to wind and water erosion of the soil and dispersion of the contaminants to adjacent areas. As a result, ecosystem and human health are put at risk via exposure through the food chain, drinking water and air.
Phytoremediation uses metal-tolerant plants (i.e. metallophytes) to extract or stabilize metals in the soils of contaminated sites. Plant establishment for phytoremediation may be limited and delayed when soil contamination is severe. Reduction of contaminant concentrations in the surface soil is a common but not always effective means of enhancing plant establishment on polluted sites. The understanding of plant responses to various concentrations of metals in soils and their metal accumulation characteristics is crucial to selection of the metallophyte plants most appropriate for specific phytoremediation purposes. The design and use of micron-size hydrogel particles that have the capacity to bind toxic soluble metals and supply water offers an effective means to enhance plant establishment.
A literature review identified new key parameters for the characterization of plant responses to metals in soil. Based on metal transporter kinetic parameters, a conceptual framework of plant metal uptake in relation to plant available metal concentration in soil was developed, and a new associated terminology for metallophytes is proposed, i.e. metal tolerators. The framework applies to all plant parts and plant available metal concentrations in soils, and was validated using independent datasets from field surveys of Queensland native plant species and the literature. This new framework may be a useful tool for selecting suitable metal tolerators for specific phytoremediation purposes, and may be also applied to non-metal elements or ions.
The thesis also examined four Australian native plant species, Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla for their tolerances to different concentrations of arsenic (As(V), 13.34–667.36 µM), copper (Cu2+, 0.5-200 µM), zinc (Zn2+, 9-500 µM), manganese (Mn2+, 8-10240 µM) and lead (Pb2+, 240-9600 µM) in single solutions during germination in controlled laboratory conditions. Metal/loid tolerance indicators used were maximum germination percentage (Gmax), mean germination time (MGT), radicle and shoot tolerance indexes (RTI & STI). Radicle tolerance index was the most sensitive indicator of metal ii tolerance in germinating seeds. All native species were highly tolerant to the metal/loids tested, however, they showed different metal toxicity thresholds and levels of tolerance based on RTI as a metal tolerance indicator during germination. Overall, all four species could be classified as metallophytes, confirming their current suitability and use for mine site rehabilitation. This work may also serve as a basis for future studies on metal/loid tolerance of other plant species during germination.
The capacity of micron-size thiol functional cross-linked acrylamide polymer hydrogel particles (X3) to facilitate germination and plant establishment of Australian native metallophyte grasses in two phytotoxic mine wastes (extremely saline and metal contaminated mine waste rock and tailings) was tested in germination and glasshouse experiments. Selection of suitable grasses to remediate both mine wastes was based on their metal tolerance characteristics established during previous germination experiments and on contaminants found in mine wastes after waste characterization.
Addition of X3 to both mine wastes significantly increased their plant-available water holding capacity (WHC) and lowered toxic soluble metal concentrations in mine waste leachates. Germination percentages and radicle elongation of both grasses selected for the germination experiments (A. lappacea and A. scabra) in wastes were significantly increased. Higher germination percentages and greater radicle development recorded in X3 amended wastes under water limited conditions suggests that X3 was able to ameliorate metal toxicity to radicles, and provide moisture, which improved the imbibition and consequent germination of the seeds.
The effectiveness of X3 to improve plant establishment and survival on both waste materials was further investigated in a short-term glasshouse experiment using A. lappacea as a model species. Addition of X3 to the top 50 mm layers of phytotoxic waste rock and tailings in pots significantly improved substrate conditions. Total soluble concentrations of metals in the top layers of both substrates were reduced immediately after X3 was applied. The conditions of the bottom layers of waste rock and tailings were also enhanced after 61 days of X3-application, with significantly reduced substrate salinity and significantly increased water retention and availability of water to plants. X3 proved to be more effective in ameliorating phytotoxic materials when combined with metallophytes. Root penetration was greatly increased in the presence of X3 particles. Improved substrate conditions through application of X3 allowed successful early establishment of A. lappacea in phytotoxic waste rock and tailings where plants failed to establish in untreated waste.
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Further research is needed to investigate whether the addition of X3 also facilitates long-term plant establishment in contaminated substrates.
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Declaration by author
This thesis is composed of my original work, and contains no material previously published or written by another person except where due reference has been made in the text. I have clearly stated the contribution by others to jointly-authored works that I have included in my thesis.
I have clearly stated the contribution of others to my thesis as a whole, including statistical assistance, survey design, data analysis, significant technical procedures, professional editorial advice, financial support and any other original research work used or reported in my thesis. The content of my thesis is the result of work I have carried out since the commencement of my higher degree by research candidature and does not include a substantial part of work that has been submitted to qualify for the award of any other degree or diploma in any university or other tertiary institution. I have clearly stated which parts of my thesis, if any, have been submitted to qualify for another award.
I acknowledge that an electronic copy of my thesis must be lodged with the University Library and, subject to the policy and procedures of The University of Queensland, the thesis be made available for research and study in accordance with the Copyright Act 1968 unless a period of embargo has been approved by the Dean of the Graduate School.
I acknowledge that copyright of all material contained in my thesis resides with the copyright holder(s) of that material. Where appropriate I have obtained copyright permission from the copyright holder to reproduce material in this thesis and have sought permission from co-authors for any jointly authored works included in the thesis.
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Publications during candidature
Peer-reviewed papers Guterres, J., Rossato, L., Doley, D., Pudmenzky, A. 2019. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters. Journal of Hazardous Materials, Vol. 364, p.449-467.
Guterres, J., Rossato, L., Doley, D., Pudmenzky, A., Bee, C., Cobena, V. 2019. Assessing germination characteristics of Australian native plant species in metal/metalloid solution. Journal of Hazardous Materials, Vol. 364C, p.173-181.
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2013. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248– 249, p.442– 450.
Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Pillai-McGarry, U., Schmidt, S. 2013. Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248– 249, p.424– 434.
Conferences proceedings Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2011. Micron- sized metal-binding hydrogel particles improve germination and root elongation of Australian metallophyte grasses (Astrebla lappacea and Austrostipa scabra) on highly polluted waste rock and tailings. Proceedings of the 32 Australasian Polymer Symposium 13-17 February 2011. Coffs Harbour, New South Wales, Australia.
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Butler, D., Brushe, J. 2011. A geobotanical survey of serpentine and metalliferous mine sites in Australia led to the discovery of new native metallophyte species and their metal-accumulation characteristics. Proceedings of the Seventh International Conference on Serpentine Ecology 12-16 June 2011, Coimbra, Portugal.
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2012. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian vi metallophyte grasses in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia, page 517-532.
Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whitakker, M., Pillai-McGarry, U., Schmidt, S. 2012. Novel metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia, page 533-550.
Oral presentation at international peer reviewed conferences Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Butler, D., Brushe, J. 2011. A geobotanical survey of serpentine and metalliferous mine sites in Australia led to the discovery of new native metallophyte species and their metal-accumulation characteristics. Presented by the candidate, International Conference of Serpentine Ecology 12-16 June 2011, Coimbra, Portugal.
Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whitakker, M., Pillai-McGarry, U., Schmidt, S. 2012. Micron-size metal-binding hydrogel particles alleviate soil toxicity and facilitate germination, root elongation and healthy early establishment of Australian metallophyte grasses in mine wastes. Joint presentation, presented by the candidate and Marie Bigot, Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia.
Publications included in this thesis
Journal Paper: Guterres, J., Rossato, L., Doley, D., Pudmenzky, A. 2019. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters. Journal of Hazardous Materials, Vol. 364, p.449-467. – incorporated as Chapter 3. Contributor Statement of contribution Guterres, J (Candidate) Conception and design (45%) Analysis and interpretation (50%) Drafting and production (65%) Rossato, L Conception and design (30%) Analysis and interpretation (25%) Drafting and production (15%)
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Doley, D Conception and design (20%) Analysis and interpretation (20%) Drafting and production (15%) Pudmenzky, A Conception and design (5%) Analysis and interpretation (5%) Drafting and production (5%)
Journal Paper: Guterres, J., Rossato, L., Doley, D., Pudmenzky, A., Bee, C., Cobena, V. 2019. Assessing germination characteristics of Australian native plant species in metal/metalloid solution. Journal of Hazardous Materials, Vol. 364C, p.173-181. – incorporated as Chapter 4. Contributor Statement of contribution Guterres, J (Candidate) Conception and design (40%) Analysis and interpretation (60%) Drafting and production (60%) Rossato, L Conception and design (25%) Analysis and interpretation (10%) Drafting and production (18%) Doley, D Conception and design (20%) Analysis and interpretation (10%) Drafting and production (15%) Pudmenzky, A Conception and design (10%) Analysis and interpretation (10%) Drafting and production (5%) Bee, C Conception and design (2.5%) Analysis and interpretation (5%) Drafting and production (1%) Cobena, V. Conception and design (2.5%) Analysis and interpretation (5%) Drafting and production (1%)
Journal Paper: Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2013. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian
viii metallophyte grasses in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248– 249, p.442– 450. – incorporated as Chapter 5. Contributor Statement of contribution Guterres, J (Candidate) Conception and design (45%) Analysis and interpretation (70%) Drafting and production (60%) Rossato, L Conception and design (20%) Analysis and interpretation (10%) Drafting and production (15%) Pudmenzky, A Conception and design (5%) Analysis and interpretation (10%) Drafting and production (10%) Doley, D Conception and design (10%) Analysis and interpretation (5%) Drafting and production (10%) Whittaker, M Conception and design (15%) Analysis and interpretation (5%) Drafting and production (4%) Schmidt, S Conception and design (5%) Analysis and interpretation (0%) Drafting and production (1%)
Conference Proceedings: Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2012. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia, p.517-532. – incorporated as Chapter 5. Contributor Statement of contribution Guterres, J (Candidate) Conception and design (45%) Analysis and interpretation (70%) Drafting and production (60%) Rossato, L Conception and design (20%) Analysis and interpretation (10%)
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Drafting and production (15%) Pudmenzky, A Conception and design (5%) Analysis and interpretation (10%) Drafting and production (10%) Doley, D Conception and design (10%) Analysis and interpretation (5%) Drafting and production (10%) Whittaker, M Conception and design (15%) Analysis and interpretation (5%) Drafting and production (4%) Schmidt, S Conception and design (5%) Analysis and interpretation (0%) Drafting and production (1%)
Journal Paper: Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Pillai-McGarry, U., Schmidt, S. 2013. Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248– 249, p.424– 434. - incorporated as Appendix IX. Contributor Statement of contribution Bigot, M Conception and design (35%) Analysis and interpretation (35%) Drafting and production (40%) Guterres, J (Candidate) Conception and design (35%) Analysis and interpretation (35%) Drafting and production (34%) Rossato, L Conception and design (10%) Analysis and interpretation (10%) Drafting and production (10%) Pudmenzky, A Conception and design (10%) Analysis and interpretation (10%) Drafting and production (5%) Doley, D Conception and design (5%) Analysis and interpretation (5%)
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Drafting and production (5%) Whittaker, M Conception and design (2.5%) Analysis and interpretation (2.5%) Drafting and production (2.5%) Pillai-McGarry, U Conception and design (2.5%) Analysis and interpretation (2.5%) Drafting and production (2.5%) Schmidt, S. Conception and design (0%) Analysis and interpretation (0%) Drafting and production (1%)
Conference Proceedings: Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whitakker, M., Pillai-McGarry, U., Schmidt, S. 2012. Novel metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia Page 533-550. – incorporated as Appendix IX. Contributor Statement of contribution Bigot, M Conception and design (35%) Analysis and interpretation (35%) Drafting and production (40%) Guterres, J (Candidate) Conception and design (35%) Analysis and interpretation (35%) Drafting and production (34%) Rossato, L Conception and design (10%) Analysis and interpretation (10%) Drafting and production (10%) Pudmenzky, A Conception and design (10%) Analysis and interpretation (10%) Drafting and production (5%) Doley, D Conception and design (5%) Analysis and interpretation (5%) Drafting and production (5%) Whittaker, M Conception and design (2.5%)
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Analysis and interpretation (2.5%) Drafting and production (2.5%) Pillai-McGarry, U Conception and design (2.5%) Analysis and interpretation (2.5%) Drafting and production (2.5%) Schmidt, S. Conception and design (0%) Analysis and interpretation (0%) Drafting and production (1%)
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Contributions by others to the thesis
Chapter 1 and 2 are written by the candidate with structural and editorial advice from the candidate’s supervisors Dr Laurence Rossato, Dr David Doley, and from Dr Alex Pudmenzky.
Chapter 3 is taken directly from a manuscript accepted and published in Journal of Hazardous Materials by Jaquelina Guterres Correia, Laurence Rossato, David Doley and Alex Pudmenzky. The manuscript was written by the candidate with editorial advice from the candidate’s supervisors. Data collection was conducted by Dr Don Butler and Ms Joy Brushe from Queensland Herbarium (Brisbane, Australia), as well as by staff from mine sites and Dr Laurence Rossato and Dr Alex Pudmenzky. Sample preparation and ICP analyses were conducted by the candidate under the supervision of Dr Laurence Rossato, Dr David Doley and Dr Alex Pudmenzky with the assistance of Mr David Appleton and Mr Steve Appleton from the School of Agriculture and Food Sciences’ analytical services at the University of Queensland. Data analysis and preparation of figures were performed by the candidate and under the supervision of the advisors Dr Laurence Rossato and Dr David Doley, and Dr Alex Pudmenzky.
Chapter 4 is taken directly from a manuscript accepted and published in Journal of Hazardous Materials by Jaquelina Guterres Correia, Laurence Rossato, David Doley, Alex Pudmenzky, Christina Bee and Victor Cobena. The manuscript was written by the candidate with editorial advice from the candidate’s supervisors. The experiment was conducted by the candidate with assistance of Ms Christina Bee (a 2009 UQ Summer Scholarship Student), Mr Victor Cobena (a 2007 UQ Master student) and under the supervision of the advisors Dr Laurence Rossato, Dr David Doley and Dr Alex Pudmenzky. The ideas and the design of the experiments were by the candidate and her supervisors. Data analysis was conducted by the candidate with the assistance of her supervisors. MATLAB statistical analysis was performed by Dr Alex Pudmenzky and experiments photographs were taken by Drs Alex Pudmenzky and Laurence Rossato.
Chapter 5 is taken directly from a manuscript accepted and published in Journal of Hazardous Materials and Conference Proceedings by Jaquelina Guterres Correia, Laurence Rossato, Alex Pudmenzky, David Doley, Michael Whittaker and Susanne Schmidt. The manuscript was written by the candidate with editorial advice from Drs Laurence Rossato, David Doley, Alex Pudmenzky and Michael Whittaker. The idea and design of the study were generated by the candidate and Drs Laurence Rossato, Alex Pudmenzky, David Doley, Michael Whittaker and Prof Susanne Schmidt. Production of metal-binding particles (X3) was funded by the University of New South Wales
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Centre for Advanced Macromolecular Design (CAMD). X3 particles were prepared by Dr Michael Whittaker and the candidate, with assistance from Dr Laurence Rossato and Dr Alex Pudmenzky. The experiment was conducted by the candidate with the assistance of Mr Craig Claremont (a 2010 UQ Summer Scholarship Student) and advice from Drs Laurence Rossato, David Doley and Alex Pudmenzky. Photographs taken during the experiment were by Drs Alex Pudmenzky and Laurence Rossato. Data analysis was conducted by the candidate with assistance from Drs Laurence Rossato, David Doley and Alex Pudmenzky and Michael Whittaker. Tailings and waste rock substrates were donated by operating and closed mine sites in Queensland (Australia).
Chapter 6 is written by the candidate with structural and editorial advice from the candidate’s supervisors Dr Laurence Rossato and Dr David Doley.
Appendix IX is taken directly from a manuscript accepted and published in Journal of Hazardous Materials and Conference Proceedings by Marie Bigot, Jaquelina Guterres Correia, Laurence Rossato, Alex Pudmenzky, David Doley, Michael Whittaker, Usha Pillai-McGarry and Susanne Schmidt. The experiment was conducted by the candidate and Ms Marie Bigot (the first author) with the assistance of Ms Christina Bee (honours student), Mrs Jenny Vu and Drs Laurence Rossato, Alex Pudmenzky, David Doley and Usha Pillai-McGarry. Production of metal-binding particles (X3) was funded by the University of New South Wales Centre for Advanced Macromolecular Design (CAMD). X3 particles were prepared by Dr Michael Whittaker. Soil psychrometer measurements and data interpretation were performed by Dr Usha Pillai-McGarry. Sample preparation and ICP analyses were conducted by the candidate and Ms Marie Bigot with the help of Ms Christina Bee and with the assistance of Mr David Appleton and Mr Stephen Appleton. Tailings and waste rock substrates were donated by operating and closed mine sites in Queensland (Australia). The manuscript was written by Marie Bigot and reviewed and edited by the candidate with editorial advice from Drs Laurence Rossato, David Doley, Alex Pudmenzky, Michael Whittaker, Usha Pillai-McGarry and Susanne Schmidt.
Statement of parts of the thesis submitted to qualify for the award of another degree
None.
Research Involving Human or Animal Subjects
No animal or human subjects were involved in this research. xiv
Acknowledgements
My PhD study has been a long journey for me (2008-2018) and my study and thesis completion owe much to the varied contributions of many people. Despite some moments of doubt and some tough times which caused me to interrupt my candidature several times, I am proud to have never given up and to have completed my PhD thesis, for myself but also for the many people who assisted me and trusted me to complete it. I have stayed true to my words.
I would like to thank my principal supervisor Dr Laurence Rossato who pushed me beyond limits I never thought possible. I am sincerely grateful for Dr Laurence Rossato’s knowledge, patience, understanding, tireless support and friendship and I truly value all her precious time and advice. I also would like to thank my co-supervisor Dr David Doley for the precious advice and continued support throughout my PhD study. I would have not completed my thesis without Dr Laurence Rossato’s and Dr David Doley’s support. I would also like to thank my informal supervisor Dr Alex Pudmenzky for all the help, continued support and advice including for my germination and glasshouse experiments, statistical analysis and photographs of experiments. I would also like to thank Dr Usha Pillai-McGarry for her help with use of pressure plate, psychrometer and penetrometer resistance, and advice on data interpretation and help on finding relevant information for my thesis and the friendship she has offered.
I would like to thank: Ms Marie Bigot for assistance with the glasshouse experiments and writing and submission of our joint published paper; Ms Christina Bee and Mr Craig Claremont (2009 and 2010 UQ Summer scholarship holders) and Mr Victor Cobena (2007 UQ Master student) for assistance with the germination experiments and data entry, Dr Michael Whittaker (UNSW, Sydney) for training in and production of particles, Dr Don Butler and Ms Joy Brush (EPA, Queensland Herbarium, Brisbane) for collecting and identifying plant samples, Mr David Appleton and Mr Stephen Appleton for their help with ICP analyses and Prof Susanne Schmidt for her very constructive feedback with my proposed research plan. I also would like to thank Dr Ben Ross (Coordinator of Higher Degrees by Research) and Mrs Myrtle Sahabandu (Research and Scholarships Office) from the UQ School of Pharmacy for all the support and assistance provided to me.
I would also like to thank Mt Morgan and Cracow mines for providing waste rock and plant and soil samples, Xstrata Technology and Commercialisation Australia for funding part of my study, the Department of Employment, Economic Development and Innovation (DEEDI Mines) for their help xv and support and AusAID for awarding me a scholarship. I also would like to thank the UQ School of Pharmacy for giving me the opportunity to be part of the School to complete my PhD study.
Finally, I would like to thank my family especially my parents, my father Manuel Guterres Correia and my mother Norberta Freitas do Rego, and all my brothers and sisters, brothers and sisters in law, cousins, nieces and nephews for their love, understanding, encouragement and continued support throughout my study. I would also like to thank all my Australian and Timorese friends for their friendships and support in any form and to the Canossian Sisters in Brisbane for all the help and support I have received during my stay in Brisbane.
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Financial support
The studies in Chapters 3 and 4 were funded by a competitive University of Queensland (Brisbane, Australia) Early Career Researcher grant awarded to Dr Laurence Rossato. The research in Chapter 5 was funded by a Pathfinder grant from UniQuest Pty Ltd (The University of Queensland, Australia). The research in Appendix IX was supported by Xstrata Technology, Commercialisation Australia and the Queensland Department of Employment, Economic Development and Innovation (DEEDI Mines).
The candidate was an AusAID Scholarship recipient from January 2008 to December 2013 and also held a University of Queensland International Fees (UQI Fees) scholarship.
Keywords
Metal transporter kinetic parameters, plant metal response, metal tolerator, metal toxicity, metal- binding hydrogel particles, soil water holding capacity, mine waste remediation
Australian and New Zealand Standard Research Classifications (ANZSRC)
ANZSRC code: 060705, Plant Physiology, 40% ANZSRC code: 100203, Bioremediation, 40% ANZSRC code: 100701, Environmental nanotechnology, 20%
Fields of Research (FoR) Classification
FoR code: 0607, Plant Biology, 40% FoR code: 1002, Environmental Biotechnology, 40% FoR code: 1007, Nanotechnology, 20%
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Table of Contents
Chapter 1. Introduction ...... 30 1.1 Background ...... 30 1.2 Research objectives...... 33
Chapter 2. Literature Review ...... 35 2.1 Metals in soils ...... 35 2.1.1 Sources of metal pollutant in soils ...... 35 2.1.2 Factors affecting mobility and bioavailability of metals ...... 38 2.1.3 Determination of metals in soils (and plants) ...... 47 2.2 Metals in plants ...... 52 2.2.1 Metal uptake by plants ...... 52 2.2.2 Metal translocation to shoots...... 53 2.2.3 Symptoms of metal toxicity in plants...... 54 2.3 Plant responses to metals ...... 60 2.3.1 Metal excluders ...... 60 2.3.2 Metal accumulators...... 61 2.3.3 Metal hyperaccumulators ...... 61 2.3.4 Metal indicators ...... 63 2.4 Tolerance mechanisms in plants ...... 63 2.4.1 Tolerance mechanisms in germinating seeds ...... 63 2.4.2 Tolerance mechanisms in established plants ...... 64 2.5 Phytoremediation ...... 68 2.5.1 Phytoextraction ...... 69 2.5.2 Phytostabilization ...... 69 2.6 Nanotechnology ...... 71 2.6.1 Nano/microparticles and their applications in agriculture ...... 71 2.6.2 Nano/microparticles and metals ...... 72 2.6.3 Nano/microparticles and the environment...... 73 2.7 Hypothesis...... 73
Chapter 3. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters ...... 75 3.1 Abstract ...... 75 3.2 Introduction ...... 76 xviii
3.3 Materials and methods ...... 80 3.3.1 Plant and soil sample collection and treatment...... 80 3.3.2 Soil analysis ...... 81 3.3.3 Plant analysis ...... 83 3.3.4 Assessing possible plant sample contamination ...... 83 3.3.5 Data analysis...... 90 3.4 Results ...... 91 3.4.1 Conceptual framework for plant metal uptake from soils ...... 91 3.4.2 Conceptual framework validation using field survey of Queensland native plant species 114 3.5 Discussion ...... 137 3.6 Conclusion ...... 139
Chapter 4. Assessing germination characteristics of Australian native plant species in metal/loid solution ...... 140 4.1 Abstract ...... 140 4.2 Introduction ...... 141 4.3 Materials and methods ...... 142 4.3.1 Plant materials ...... 142 4.3.2 Chemicals...... 143 4.3.3 Germination experiments ...... 143 4.3.4 Determination of metal tolerance indicators ...... 145 4.3.5 Statistical analysis ...... 146 4.4 Results ...... 146 4.4.1 Effect of metal/loids on metal tolerance indicators during germination ...... 146 4.4.2 Effect of metal/loids on radicle tolerance index (RTI)...... 147 4.5 Discussion ...... 153 4.6 Conclusion ...... 156
Chapter 5. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings ...... 157 5.1 Abstract ...... 158 5.2 Introduction ...... 158 5.3 Materials and methods ...... 160 5.3.1 Substrate collection and characterisation ...... 160
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5.3.2 Determination of metal-binding and water holding capacities of X3 in mine substrates under various pH conditions...... 161 5.3.3 Measurement of germination characteristics and radicle elongation of A. lappacea and A. scabra in mine substrates amended with X3 under different water regimes ...... 162 5.3.4 Determination of pH, plant available metal concentrations and cumulative evaporation in germination substrates...... 163 5.4 Data analysis ...... 163 5.5 Results ...... 163 5.5.1 Characteristics of substrates...... 163 5.5.2 Effect of X3 particles on soluble metal concentrations in mine waste leachates ...... 164 5.5.3 Water holding capacities of mine substrates amended with X3 ...... 165 5.5.4 Effect of X3 on seed germination ...... 166 5.6 Discussion ...... 173 5.7 Conclusion ...... 175
Chapter 6. General Discussion and Conclusions...... 177 6.1 Objective 1. Examine the relationship between selected Australian plant species and the properties of metal-rich or contaminated sites ...... 178 6.1.1 To develop a new conceptual framework to characterize plant responses to soil metals based on metal transporter kinetic parameters, and validate this model using field survey of Queensland native plant species and literature data...... 178 6.1.2 To examine the effect of arsenic (AsV), copper (Cu), zinc (Zn), manganese (Mn) and lead (Pb) on the germination characteristics of potential native metallophytes including Astrebla lappacea, Austrostipa scabra, Themeda australis and Acacia harpophylla in the absence of metal- binding particles...... 180 6.2 Objective 2. Modification of conditions to enhance plant establishment on contaminated substrates ...... 183 6.2.1 To assess the potential of metal-binding particles to sequester metals in contaminated substrates and to increase substrate water holding capacity...... 183 6.2.2 To investigate the effect of metal-binding particles on germination of selected native metallophyte grasses (A. lappacea and A. scabra as model species) on contaminated substrates. 184 6.2.3 To examine the effect of metal-binding particles on plant establishment and growth of the native metallophyte grass A. lappacea on contaminated substrates under controlled conditions in a short-term glasshouse trial...... 185 6.3 Concluding remarks and future research ...... 187
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References ……………………………………………………………………………………… 190 Appendix I ….…………………………………………………………………………………… 250 Appendix II ……………………………………………………………………………………… 252 Appendix III ...…………………………………………………………………………………... 253 Appendix IV …………………………………………………………………………………….. 254 Appendix V ……………………………………………………………………………………... 255 Appendix VI …………………………………………………………………………………….. 256 Appendix VII ……………………………………………………………………………………. 257 Appendix VIII …………………………………………………………………………………… 260 Appendix IX …………………………………………………………………………………….. 261 Appendix X ……………………………………………………………………………………... 287 Appendix XI …………………………………………………………………………………….. 287 Appendix XII …………………………………………………………………………………… 287 Appendix XIII …………………………………………………………………………………... 287
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List of Figures
Figure 2-1 Serpentine rocks and soils (shaded) in Central Queensland (Source: Batianoff et al., 1990)...... 36 Figure 2-2 Modelled adsorption of certain trace metals onto hydrous ferric oxide (Source: Adriano, 2001)...... 39 Figure 2-3 Average germination percentage of six varieties of wheat growing in different concentrations of arsenite () and arsenate () (Source: Liu et al., 2005)...... 45 Figure 2-4 Different approaches applied in general to the determination of metals in soils (% values shown are approximate as they are metal and matrix dependent) (Source: Rao et al., 2008). .. 48 Figure 2-5 Simplified structure of the root, showing apoplastic (left) and symplastic (right) pathways of radial metal transport and exchanges between apoplast and symplast (bottom) between the root surface and the stele (Source: McLaughlin, 2002)...... 53 Figure 2-6 Effect of increasing concentrations of arsenate on early root growth of A. lappacea (Source: Guterres et al., 2019)...... 56 Figure 2-7 Copper toxicity (reddening of leaflet and stunted growth) in (a) Eucalyptus crebra (b) Acacia holosericea (Source: Reichman, 2001)...... 58 Figure 2-8 Zinc toxicity in Melaleuca leucadendra (a) Leaf curling and wilting (b) Red colouration of the old leaves (Source: Reichman, 2001)...... 58 Figure 2-9 Aluminium toxicity in Triticum aestivum L. (a) Stubby appearance and brownish colour of roots (b) Leaves showing yellowing/chlorosis along the margin and formation of necrotic indentations (Source: International Maize and Wheat Improvement Centre, 2010)...... 59 Figure 2-10 Relationships between plant tissue metal concentration and total metal concentrations in soil for the four different types of plant responses to metals (Source: Adriano, 2001)...... 61 Figure 2-11 Surface view of a seed of Grevillea exul var. rubiginosa; a) seed coat; b) position of the embryo; c) endosperm area (Source: Leon et al., 2005)...... 64 Figure 2-12 SEM images and point analysis in the seed coat of a germinating seed of G. exul var. rubiginosa treated with 500 mg Ni L-1; (A) Longitudinal section of a seed: a) endosperm; b) seed coat with a hypodermal cell layer incrusted by crystals; (B) Detail of the inset area in A indicating crystal locations: a) outer layer of the seed coat; b) crystal in the seed coat; c) sclereid cell in the seed coat; d) inner layer of the seed coat (Source: Leon et al., 2005)...... 64 Figure 3-1 Conceptual model of changes in metal uptake into plants, integrated over time, as a function of low and high plant available metal concentrations in the substrate. Experimental (EXP) plants grown in controlled laboratory conditions in solution are either deprived (grown without metal), initiating a constitutive high-affinity transport system (CHATS) at low xxii
experimental metal concentrations and both constitutive high- and low-affinity transport systems (CLATS+CHATS) at high experimental metal concentrations, (brown lower curve), or metal-supplied (grown in the presence of metal), initiating an inducible high-affinity transport system (IHATS) at low experimental metal concentrations and both inducible high- and low-affinity transport systems (ILATS+IHATS) at high experimental metal concentrations, (blue middle curve). Uptake attributed to the CHATS or IHATS follows
Michaelis-Menten kinetics with parameters Vmax and Km (the saturable component). When substrate metal concentrations reach the transition concentration (shown by the vertical line above the asterisk in the figure), low-affinity transport systems (CLATS and ILATS) whereby metal uptake increases linearly with increasing substrate concentration (the linear component), dominate total metal uptake (CHATS+CLATS and IHATS+ILATS). The linear portions of the uptake curves are identified by the slope and intercept (crosses at zero-concentration) for each transport system. For plants collected from the field, uptake will be the combined result of both constitutive and inducible uptake (FIELD HATS at substrate plant available metal concentrations below the transition concentration and FIELD HATS+LATS at greater substrate plant available metal concentrations, red upper curves). At very high external metal concentrations, mechanisms of plant metal resistance or tolerance break down and unrestricted transport occurs causing plant death...... 91 Figure 4-1 Radicle elongation of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days of exposure to various concentrations of As(V) solutions. Arrows indicate emergent radicle...... 150 Figure 4-2 Radicle elongation of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days of exposure to various concentrations of Pb solutions. Arrows indicate emergent radicle...... 152 Figure 5-1 Cumulative water loss from sand, waste rock and tailings for A. lappacea (a and b) and A. scabra (c and d) in Petri dishes maintained at field capacity or water limited conditions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean (n=5). 168 Figure 5-2 Radicle elongation of A. lappacea after 4 days and A. scabra after 3 days at field capacity and in water limited regime on unamended and X3 amended waste rock (18.4% DW) and tailings (3.2% DW). Arrows indicate emergent radicle...... 172
xxiii
List of Tables
Table 2-1 Categories of metal occurrence in soil, typical sequential extraction methods and typical percentage of total metal released (Compiled from Rao et al., 2008, Zimmerman and Weindorf, 2010, Kim et al., 2015)...... 49 Table 2-2 Examples of extractants used for assessing plant-available trace elements concentration in soil (from numbered sources, 1-27 cited at the bottom of the Table)...... 51 Table 2-3 Metal toxicity limits...... 55 Table 2-4 Threshold foliar metal dry weight concentrations in non-tolerant plants (agricultural crops and pastures) and metal accumulators and hyperaccumulators from literature...... 62 Table 3-1 Summary of plant and soil sample identification used in this study...... 81 Table 3-2 Estimated total concentrations of selected metals in soil fine material and plant dry matter (mg.kg-1) and soil/plant ratios (Mitchell, 1960)...... 87 Table 3-3 Lowest observable adverse effects concentrations (LOAECs) and associated thresholds used for monitoring possible soil contamination of plant samples. Values are not given for very sensitive or highly tolerant plant species. Concentrations of Cr, Al and Fe in soil refer to total concentrations of these metals in soil. “Plant” in soil/plant ratios refers to above ground plant parts...... 89 Table 3-4 Transporter kinetic parameters of essential and non-essential elements in plants, extracted from the literature. Values are given as the mean ± standard errors (when available from the articles). The number of replicates varied from 2 to 27...... 95 Table 3-5 Metal concentrations in soil at each plant species collection site. Values are given as the mean ± standard errors (n=3)...... 115 Table 3-6 LATS metal uptake parameters derived from field collected plant species. Parameter values are given as the mean for each species across all sites...... 122 Table 3-7 HATS metal uptake parameters derived from field collected plant species. Parameter values are given as the mean for each species across all sites...... 127 Table 3-8 Total metal concentration of each metal analysed via ICP in plant parts growing in mine sites and serpentine areas. Values are given as the mean ± standard errors (n=3)...... 130 Table 4-1 Metal/loid concentrations tested on seed germination of native plant species Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla...... 144 Table 4-2 Metal/loid tolerance levels of Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla during germination on As(V)(13.34-667.36 μM), Cu2+(0.5- 200 μM), Zn2+(9-500 μM), Mn2+(8-10240 μM) and Pb2+(240-9600 μM), based on maximum
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germination percentage (Gmax), mean germination time (MGT), radicle and shoot tolerance indexes (RTI & STI) as tolerance indicators...... 147 Table 4-3 Radicle tolerance index (RTI) of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days or Acacia harpophylla after 9 hours on various concentrations of As(V), Cu, Zn, Mn and Pb solutions...... 148 Table 5-1 Characterisation of waste rock, tailings and sand (control). Results are given as the mean ± standard error (SE) (n=3)...... 164 Table 5-2 Soluble concentrations of metals in leachate of waste rock at different percentages of X3
and under various pH conditions in DIW and 0.01 M CaCl2. Data are presented as the mean ± standard error (SE) (n=3); nt stands for “not tested”...... 165 Table 5-3 Water holding capacities of waste rock and tailings amended with different percentages
of X3 and under various pH conditions in DIW and 0.01 M CaCl2. Data are presented as the mean ± standard error (n=3); nt stands for “not tested”...... 166
Table 5-4 pH and plant available metal concentrations (0.01 M CaCl2) in unamended and X3 amended (18.4% DW) waste rock used for the germination of A. lappacea and A. scabra under two different water regimes. Data are presented as the mean ± standard error (SE) (n=3). .... 167 Table 5-5 Maximum germination percentages (MaxGerm %) and mean germination time (MGT) of A. lappacea and A. scabra on unamended and X3 amended waste rock (18.4% DW of X3) and tailings (3.2% DW of X3) under different water regimes. Results are given as the mean ± standard error (SE) (n=5). Within each substrate, treatments with the same letter do not differ significantly (P>0.05); asterisks indicate values based on one replicate only...... 170 Table 5-6 Radicle tolerance index (RTI) of A. lappacea and A. scabra in unamended and X3 amended waste rock (18.4% DW of X3) and tailings (3.2% DW of X3). Results are given as the mean ± standard error (SE) (n=5). Within each substrate, treatments followed by the same letter do not differ significantly (P>0.05); asterisks indicate values based on one replicate only...... 173
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List of Abbreviations
ANOVA Analysis of variance
CaCl2 Calcium Chloride CEC Cation Exchange Capacity CHATS Constitutive High Affinity Transport System CLATS Constitutive Low Affinity Transport System CSIRO Commonwealth Scientific and Industrial Research Organisation DEEDI Department of Employment, Economic Development and Innovation DW Dry weight EDTA Ethylenediaminetetraacetic acid EPA Environmental Protection Agency
Gmax Maximum Germination Percentage GPS Geographic Information System GSH Glutathione HATS High Affinity Transport System HCl Hydrochloric acid HF Hydroflouric acid ICP-MS Inductively coupled plasma mass spectrometry ICP-OES Inductively coupled plasma optical emission spectroscopy IHATS Inducible High Affinity Transport System ILATS Inducible Low Affinity Transport System IRT Iron Transport LATS Low Affinity Transport System LOAEC Lowest observable adverse effects concentration LPSDP Leading Practice Sustainable Development Program MaxGerm Maximum Germination MGT Mean Germination Time MWCNT Multiwalled carbon nanotubes NIST National Institute Standard and Technology NRAMP Natural Resistance Associated Macrophage Protein PAM Polyacrylamide PAW Plant available water PCs Phytochelatins PC2 Physical containment level 2 xxvi
PPTR tube Polypropylene Transparent tube RTI Root Tolerance Index SAR Specific absorption rate SDIW Sterilised Deionised water SE Standard error SRM Standard Reference Material STI Shoot Tolerance Index SWCNT Single walled carbon nanotubes TDI Triple deionised water WC Water content WHC Water holding capacity
Ψo Osmotic potential
Ψm Matric potential
ΨT Total water potential
xxvii
Thesis Overview
The thesis consists of seven chapters with the following outlines:
Chapter 1 introduces the thesis topic and presents the problems related to the topic. A list of research objectives is also presented.
Chapter 2 presents a comprehensive review of the scientific literature on the topic. This includes the occurrence and movement of metals in soils and plants, soil-metal-plant relationships, metal toxicity in plants, and application of current soil remediation technologies. The chapter also includes research hypotheses.
Chapter 3 focuses on the first aim of Objective 1 of the thesis which is ‘to develop a new conceptual framework to characterize plant responses to soil metals based on metal transporter kinetic parameters’. An extensive literature review was conducted on plant responses to metals in solutions/soils. The new conceptual framework was developed and was validated using data from a field survey of Queensland native plant species and literature data.
Chapter 4 assesses the second aim of Objective 1 of the thesis which is ‘to examine the effect of arsenic (As(V)), copper (Cu2+), zinc (Zn2+), manganese (Mn2+) and lead (Pb2+) on the germination characteristics of the Australian native species Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla in the absence of micron-sized metal-binding particles’. The species’ tolerances to different concentrations of As(V), Cu2+, Zn2+, Mn2+ and Pb2+ were tested and their metal toxicity thresholds established during germination using four metal tolerance indicators (namely germination percentages, mean germination time, root and shoot tolerance indexes).
Chapter 5 evaluates the first two aims of Objective 2 of the thesis which are ‘to assess the potential of micron-sized metal-binding particles to bind metals in contaminated substrates and to increase substrate water holding capacity’ and ‘to investigate the effect of metal-binding particles on germination of Australian native plant species (A. lappacea and A. scabra as model species based on experimental results of Chapter 4) on contaminated substrates’. Phytotoxic waste rock and tailings collected from Australian mine sites were used as germination media. The effect of X3 addition on the water holding capacity (WHC), concentrations of toxic soluble metals and germination percentages and radicle elongation of both A. lappacea and A. scabra in both mine wastes were investigated.
Chapter 6 discusses significance of the key research findings in the light of the past and current
xxviii literature and identifies knowledge gaps and possible future research to address these gaps.
Appendix IX examines the third aim of Objective 2 of the thesis which is ‘to examine the effect of metal-binding particles on plant establishment and growth of A. lappacea on contaminated substrates under controlled conditions in a glasshouse trial (pot trial)’. Waste rock and tailings used in this study were collected from the same mine sites as in Chapter 5. A short-term glasshouse experiment was conducted to further investigate the effectiveness of X3 to improve plant establishment and survival of A. lappacea on both waste materials.
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Chapter 1. Introduction
1.1 Background Metals and metalloids such as arsenic (As), lead (Pb), copper (Cu), manganese (Mn), zinc (Zn), cadmium (Cd) and aluminium (Al), (collectively described here as metals) occur naturally in rocks and soils and, with the exception of As, Pb and Cd, they are mostly present in forms and at concentrations that are not toxic to humans or other living organisms (Kabata-Pendias, 2001, Alloway, 2013). Anthropogenic activities, such as mining and agriculture, have increased the release of toxic bioavailable metals into ecosystems, leading to major environmental and health problems (Nedelkoska and Doran, 2000, Bondada and Ma, 2003, Pilon-Smits, 2005, Lottermoser, 2010a, Alloway, 2013, Bolan et al., 2014).
Soils associated with mining by-products such as waste rock (material rejected during winning of the mineral ore) and tailings (usually fine-grained residues from mineral extraction processes) (Chaturvedi et al., 2015, Muñoz et al., 2016) may have high concentrations of contaminants that prevent the growth of protective layers of vegetation, leading to wind and water erosion of the soil and dispersion of the contaminants to adjacent aquifers and other areas (Alvarez et al., 2003, Mendez and Maier, 2008a). The contaminants may be retained by the soil, leached, or absorbed by vegetation (Alvarez et al., 2003), from which they may enter the food chain and put human health at risk (Gall et al., 2015).
In 2004, there were potentially 160,000 to 200,000 metal contaminated sites in Australia and 50% of them were polluted with As (CSIRO, 2004a). The majority of these sites were located in rural areas (for example, small sheep and cattle dips) or on former mining sites and fewer than 5% of them had been treated with different bioremediation methods (Naidu et al., 2003a). The number of contaminated sites in Australia may have increased since 2004 as they are difficult to revegetate while mining activities are increasing (Unger et al., 2012).
Management and remediation of sites contaminated with metals and organic compounds cost Australian industry billions of dollars per year. For example, in Queensland alone, around $189 million per year was spent on managing acid sulphate soils (Sutherland and Powell, 2000, National Working Party on Acid Sulfate Soils, 2000, Powell and Martens, 2005, Michael, 2013). The numerous existing methods, such as physico-chemical extraction of metals by enhanced leaching (Li et al., 2011a) or in situ immobilization of contaminants (e.g. by vitrification) are mostly drastic and expensive (Reddy, 2010, Komárek et al., 2013). They remove all the biological activity from
30 the soil and affect its physical structure, but they may be suitable only for small sites of valuable redevelopment land (Baker et al., 1994a). There are no acceptable techniques for remediation of large rural contaminated sites which retain soil fertility after metal removal. Current remediation methods are neither sustainable, nor sufficient to address the risks posed by contaminants (CSIRO, 2004b, Lu et al., 2017a). For example, one of the soil remediation methods most commonly applied until recently, excavation to landfill, is environmentally disruptive, costly and no longer acceptable, especially in the case of extensive contaminated sites.
A promising cost effective and environmentally benign in situ biological remediation technology involving plants, known as phytoremediation, could be utilized for environmentally sensitive contaminated sites (Hartley and Lepp, 2008). Phytoremediation is defined as the use of plants and heterotrophic soil microbes to remove pollutants from the environment or to render them harmless (Salt et al., 1995). The establishment of plants in mine tailings improves conditions for the heterotrophic microbial community and thereby may promote further plant growth, including plant species less tolerant to metals (Whiting et al., 2004), and contribute to metal stabilization (Mendez and Maier, 2008a).
There are two strategies of phytoremediation of metal contaminated sites. Phytoextraction is the use of metal-accumulating plants to transfer metals from contaminated soils and accumulate them in harvestable plant parts, such as shoots and roots, although shoots are more easily harvestable than roots, especially from soils (Salt et al., 1995, Sessitsch et al., 2013, Lu et al., 2017a). This technique involves plants which are native to metalliferous areas or tolerant to soil metal contaminants, and can accumulate high concentrations of metals (w/w) in their upper parts (hyperaccumulators) e.g. 1% for Mn and Zn, 0.1% for Ni, Al, Co, Cr, Cu, Pb, As, 0.01% for Cd and Se, and 0.001% for Hg (Baker and Brooks, 1989, Baker et al., 1994a, Pilon-Smits, 2005, Krämer, 2010). The major advantage of this technology is that contaminants are transferred from the soil to plant tissues without destroying the soil structure or fertility. Phytoextraction is most effective for the remediation of diffusely contaminated areas where metals occur at fairly low to moderate concentrations (Ghosh and Singh, 2005). However, it is a slow process, requiring 1 to 20 years (Kumar et al., 1995, Blaylock and Huang, 2000, Shahid et al., 2012) even with the best known metal hyperaccumulator, alpine penny-cress (Thlaspi caerulescens) (Baker et al., 1994b), and plant parts used in phytoextraction require harvesting and disposal (Mendez and Maier, 2008a, Bhargava et al., 2012, Deng et al., 2016).
The second strategy is phytostabilization, which is the use of metal tolerant plants to stabilize toxic metals in soils and reduce the risk of erosion and contaminant dispersion via wind and percolated 31 water (Salt et al., 1995, Meier et al., 2012, Bolan et al., 2014). This technique involves the sequestration of metals within the rhizosphere through absorption and accumulation by roots, adsorption onto roots or precipitation within the root zone (Cunningham et al., 1995, Susarla et al., 2002, Weyens et al., 2009). As a result, the bioavailability of metals to wildlife and human exposure are reduced to acceptable concentrations (McCutcheon and Schnoor, 2003, Mendez and Maier, 2008a, Shahid et al., 2014). Phytostabilization does not require harvesting of plants and is highly suited for treating large areas with elevated surface metal contamination (e.g. mine tailings) where quick immobilisation is required to reduce risks of erosion and uncontrolled transfers of metals into the environment (Testiati et al., 2013). However, because the pollutants are not removed from the soils, continued adaptive monitoring and management are needed (Ghosh and Singh, 2005, Epelde et al., 2014). The processes involved in plant establishment in elevated metal environments are identical for species being considered for either phytoextraction or phytostabilization. Consequently, the general objective of phytoremediation will be adopted in most circumstances throughout this thesis.
Plant species native to the area of concern are commonly targeted for phytoremediation (Newman et al., 1998, Yoon et al., 2006, Gaskin et al., 2008, Barrutia et al., 2011, Tongway and Ludwig, 2011, Testiati et al., 2013). Their advantages over non-native plants are: (1) they have developed survival mechanisms appropriate to the local climates of the arid and semiarid environments of Australia, and are tolerant to drought, salinity and low soil fertility (Mendez and Maier, 2008a); (2) the use of native plants prevents the introduction of non-native and potentially invasive species that may lead to decreased regional plant biodiversity (Keeling and Werren, 2005, Barrutia et al., 2011), even though exotic species adapted to low resource environments (Funk, 2013) or rapid environmental change (Diez et al., 2012) may become important and potentially toxic components of altered ecosystems (Lottermoser, 2011); (3) native grasses can often achieve rapid ground cover and wind erosion reduction (Williams and Currey, 2002) and; (4) native shrubs and trees offer a broad canopy cover and establish deeper root systems that prevent erosion over the longer term (Belsky et al., 1989, Tiedemann and Klemmedson, 2004, Tongway and Ludwig, 2011).
To date, only a few plant species have been used in phytoremediation, and these are mostly hyperaccumulators which can survive in metal-polluted soils, and in which metals are extracted and translocated from roots and concentrated in the shoots (Tu et al., 2002, Wang et al., 2002, Krämer, 2010, Sarma, 2011). The best known hyperaccumulators are mainly European species mostly belonging to the Brassicaceae family (Reeves and Brooks, 1983, Reeves, 1988). Australia has a unique flora with many potential metallophyte species, including 14 identified hyperaccumulators
32
(Pudmenzky et al., 2009) but there have been few studies of their use for phytoremediation. In order to utilize Australian native plant species for phytoremediation, an understanding is required of plant growth, responses, toxicity thresholds and tolerance mechanisms to metal/loids, and the soil or substrate characteristics affecting plant responses to metals (Reichman, 2002, Krämer, 2010).
It is important to note that phytoremediation may take time before it becomes effective and initial plant establishment might be limited because of the high concentrations of contaminants present in the surface soils (Alloway, 2013). In order to improve the effectiveness of phytoremediation on highly contaminated sites, it may be necessary to ameliorate the soil prior to revegetation by dilution or sequestration of the toxic components (Pilon-Smits, 2005, Rossato et al., 2011). The testing of novel amelioration techniques was relevant to this thesis.
Nanotechnology could be used for improving plant establishment in stressful sites. Nanomaterials are defined as those with at least two dimensions between 1 and 100nm (Colvin, 2003, Dionysiou, 2004, Sun et al., 2006, Klaine et al., 2008). The use of nanomaterials for commercial purposes is increasing (Lin and Xing, 2007) and its impacts are increasingly evident in all areas of science and technology, including environmental science and management (Krämer, 2010). Applications of nanotechnologies include pollutant sensing and detecting, pollution prevention, and site treatment and remediation (Khin et al., 2012). Of these applications, the latter group has undergone the most growth in recent years. For example, nanotechnology has already been used for contaminant removal or destruction (Bidar et al., 2007, Klimkova et al., 2011, Alqudami et al., 2012, Shah et al.,
2016) such as removal of Cr(VI) by Fe3O4 nanoparticles (Tang and Lo, 2013) and Cr(VI) and Cu(II) by green synthesized iron based nanoparticles (Weng et al., 2016). The effectiveness of xanthate nanoparticles in selectively removing heavy metals (alkali metals have only a low affinity for binding) from water to concentrations below 1 ppm is also another example of successful contaminant removal using nanotechnology (Chaudhari and Tare, 1999, Bell et al., 2006, Chang et al., 2007, Xu et al., 2012).
1.2 Research objectives The general objective of this thesis is to: (1) examine the relationship between selected Australian plant species and the properties of metal-rich or contaminated sites, and (2) test the contributions of synthetic designer metal-binding particles and these species to site remediation. The two components of this objective were investigated by considering: (1) Plant properties under metal conditions: The aims were to:
33
(a) develop a new conceptual framework to characterize plant responses to soil metals based on metal transporter kinetic parameters, and validate this model using field survey of Queensland native plant species and literature data. (b) examine the effect of arsenic (As(V)), copper (Cu), zinc (Zn), manganese (Mn) and lead (Pb) on the germination characteristics of potential native metallophytes, including Astrebla lappacea, Themeda australis, Austrostipa scabra, and Acacia harphophylla in the absence of metal-binding particles. (2) Modification of conditions to enhance plant establishment on contaminated substrates: The aims were to: (a) assess the potential of metal-binding particles to sequester metals in contaminated substrates and to increase substrate water holding capacity. (b) investigate the effect of metal-binding particles on germination of selected native metallophyte grasses (A. lappacea and A. scabra as model species) on contaminated substrates. (c) examine the effect of metal-binding particles on plant establishment and growth of the native metallophyte grass A. lappacea on contaminated substrates under controlled conditions in a short-term glasshouse trial. Since this research was conducted jointly with another student, corresponding results have been placed into Appendix IX of the thesis.
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Chapter 2. Literature Review
2.1 Metals in soils 2.1.1 Sources of metal pollutant in soils Metals in soils originate from parental materials and anthropogenic activities such as mining (Baker, 1987, Naidu et al., 2003b, Lottermoser, 2010a, Selinus, 2013), agriculture, energy production, microelectronics and waste/scrap disposal (Adriano, 2001, Tsydenova and Bengtsson, 2011, Bigum et al., 2012).
2.1.1.1 Natural metalliferous soils (serpentine soils/ultramafic rocks) Metals accumulate locally in soils as a result of in situ weathering of rock minerals at or near the soil surface (Ross, 1994, Burak et al., 2010, Xu et al., 2014, Alloway, 2013). Weathering is the process of dissolution, hydration, hydrolysis, oxidation, reduction, and carbonatization of rocks (Kabata-Pendias, 2001) which leads to the development of natural metalliferous soils. Metals released from parental materials are not in available forms but human intervention may produce more readily absorbed forms in which metals become toxic to plants if present in excessive amount (Wong, 1986, Lottermoser, 2010a, Plumlee and Morman, 2011, Crichton, 2017).
Serpentine or ultramafic soils are known as naturally occurring metal-rich environments. These soils originate from the weathering and pedogenesis of rocks containing ferromagnesian minerals, particularly serpentinite (Pal et al., 2005, Bani et al., 2014). Serpentine soils often display extreme edaphic conditions and support endemic and specialised metalliferous floras (metallophytes), with occasional occurrences of metal hyperaccumulators, the most widely known of which are Ni hyperaccumulator plants (Batianoff et al., 1990, Pal et al., 2005, Krämer, 2010, Callahan et al., 2012).
In Australia, serpentine soils exist mostly in Central Queensland, Western Australia and New South Wales (Bhatia et al., 2002). Of those areas, the former (Figure 2-1) represents the largest single ultramafic region in Australia with 1000 km2 area of serpentine landscape (Batianoff et al., 2000). Batianoff et al. (1990) carried out a survey in the serpentine area of Rockhampton (Central Queensland) and found Stackhousia tryonii Bailey, a nickel hyperaccumulator, with concentration of this element reaching 1-2% of the dry weight in the leaves and 0.1-1% in other plant parts. A later survey (Burge and Barker, 2010) found between 0.25 and 4.1% Ni in leaves of S. tryonii, but no evidence of hyperaccumulation in related species or genera, suggesting that the ability to hyperaccumulate nickel had occurred only once during the evolution of S. tryonii.
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Serpentine soils usually have a pH ranging from 4.5 to 6.7 (Spence and Millar, 1963) and contain high concentrations of heavy metals such as Fe, Cr, Ni, Co and Mn (Brady et al., 2005, Pal et al., 2005, Frizzi et al., 2017); and low essential plant nutrient concentrations (i.e. N, P, K and Mo) (Spence and Millar, 1963, Kazakou et al., 2008). Although some studies have reported that some serpentine soils have low Mg contens (Birrell and Wright, 1945, Spence and Millar, 1963), they normally have high Mg, low Ca contents (significantly lower Ca concentrations compared to surrounding areas), and very low calcium-to-magnesium ratios which may reduce nutrient uptake and increase Mg and Ni toxicity (Brady et al., 2005). High soil Ni and Cr concentrations have been reported to be major cause of reduced crop growth and metal toxicity symptoms on serpentine soils (Spence and Millar, 1963, Brady et al., 2005). These soils also have high concentrations of cobalt which potentially can be toxic to plants (Baker et al., 2000).
Figure 2-1 Serpentine rocks and soils (shaded) in Central Queensland (Source: Batianoff et al., 1990).
The physical conditions of ultramafic soils are also unfavourable for plant growth. The soils normally have low silt and clay contents and consequently low moisture holding capacities at intermediate water contents which result in low nutrient concentrations (Brady et al., 2005). They are often infertile and support low vegetation densities compared to the surrounding areas; consequently, erosion increases and the lack of cover leads to higher soil temperatures (Kruckeberg, 2002). Therefore, plants found in these regions are likely to tolerate drought conditions in addition to the serpentine chemical attributes (Brady et al., 2005). The survival of many metallophytes in environments with high metal availability may create a niche that is protected from other plant 36 species that are more competitive on non-toxic soils (Jaffré, 1992, Morat, 1993), but in some situations species niches may be associated more closely with metal tolerance than with comparative biomass production (Adamidis et al., 2014). It has also been shown that metal accumulation in the foliage may allow hyperaccumulator species to evade predators such as herbivores (caterpillars), fungi and bacteria as accumulated metals are acutely and directly toxic (Boyd and Martens, 1998). This “defense against pathogen/predator” scenario is believed to be one of the most plausible drivers for the evolution of metal hyperaccumulation including the interaction of this trait with other defense types (Baker et al., 2000, Horger et al., 2013), although recent work suggests that metal hyperaccumulation and tolerance traits might have an energetic cost and an impact on species fitness (Hanikenne and Nouet, 2011).
2.1.1.2 Human-induced metalliferous soils Mining is considered to be a major cause of metal contamination in soils and is the major source of transportable contaminants (Naidu et al., 2003a, Lottermoser, 2010a, Selinus, 2013, Xiao et al., 2017). There is a wide range of contaminated sites around the world resulting from mining activities. In 1995, it was estimated that on a yearly basis more than 700 million kg of metals in mine tailings from either inactive or abandoned mine sites were disposed on land in arid and semiarid regions throughout the world, including Australia (Mendez and Maier, 2008a). There are over 80,000-100,000 sites contaminated with heavy metals in Australia today, maybe twice that number if old arsenic sheep dips are included in the count (CSIRO, 2004a). Approximately 20,000 of these contaminated sites were recorded in Queensland in 2008 (Source: Environmental Management Register, Queensland EPA).
Ross (1994) and Lottermoser (2010a) identified several sources of toxic metals from metalliferous mining and smelting: Spoil heaps and dry-deposited tailings – contamination through weathering, wind erosion (As, Cd, Hg, Pb) Fluvially dispersed tailings – deposited on soil during flooding, river sediment deposition, etc. (As, Cd, Hg, Pb) Transported ore components - blown from conveyances onto soil (As, Cd, Hg, Pb) Smelting – contamination because of stack discharges and wind-blown dust, aerosols from stockpiles (As, Cd, Hg, Pb, Sb, Se, Al) Iron and steel production – stack emissions and fugitive dusts (Cu, Ni, Pb) Metal finishing - (Zn, Cu, Ni, Cr, Cd) Aluminium production (from bauxite ore) - dust emission
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Besides metal pollutants from mining and metallurgical industries, other important sources of metal pollution in soils are the semi-conductor industry, atmospheric pollution from motor vehicles, the combustion of fossil fuels, agricultural fertilisers and pesticides and organic manures, and the disposal of urban and industrial wastes leading to soil pollution from the deposition of aerosol particles emitted by the incineration of metal-containing materials (Alloway, 1990a, Alloway, 2013, Hu et al., 2013, Selinus, 2013, Oun et al., 2014).
Arsenic and lead are among the most commonly encountered metals in polluted sites and potentially the most toxic for plants, animals, human health and the environment (Huang et al., 1997, Naidu et al., 2003a, Hartley and Lepp, 2008, Lottermoser, 2010a). Their concentrations in soils are likely to increase on a global scale, with increasing mining, industrial and agricultural activities over time (Kabata-Pendias, 2001). Aluminium is the third most abundant metal in the Earth’s crust. It is released from soil minerals into the soil solution and becomes more available and toxic at pH below 5.5, which is the soil pH for 30% of arable lands (Ryan and Delhaize, 2010). Aluminium toxicity is therefore a significant limitation to plant survival and production (Ryan and Delhaize, 2010, Yang et al., 2013a).
2.1.2 Factors affecting mobility and bioavailability of metals Soil is a complex heterogeneous medium comprising mineral and organic solids, aqueous and gaseous components. It is a source of metals, a sink for metal contaminants (Alloway, 1990b, Selinus, 2013, Su, 2014) and also a physical, biological, and chemical filter (Adriano, 2001). Metals are categorized as persistent environmental pollutants since they cannot be rendered harmless by chemical or biological remediation processes (Say et al., 2002, Yoo et al., 2017). In soil, metals can be involved in a series of complex chemical and biological interactions including oxidation- reduction, precipitation and dissolution, volatilization, and surface and solution phase complexation (Selim and Amacher, 2001, Guala et al., 2010). They are not decomposed but rather only transformed, which affects their mobility in soils and transfer to plants and other organisms (Jjemba, 2005, Selinus, 2013).
Metals in soils can be categorised according to their chemical associations with different mineralogical fractions (Kadukova et al., 2008, Zimmerman and Weindorf, 2010): a) in soil solution: as free metal ions and soluble metal complexes b) adsorbed to inorganic soil constituents at ion exchange sites c) bound to soil organic matter d) precipitated as oxides, hydroxides, and carbonates e) embedded in the crystal structure of the silicate minerals. 38
Bioavailability of metals in soils is defined as the extent to which a metal present in a possible source is available for uptake (Ross, 1994, Jjemba, 2005, Almendras et al., 2007, Kim et al., 2015). It is an important factor determining metal toxicity to plants (Kabata-Pendias, 2001, Naidu et al., 2003b, Selinus, 2013). Metals in fractions (a) and (b) above are widely accepted as readily available for plant uptake, while some portions of fractions (c) and (d) may also be exchanged with plants (Harmsen, 2007, Kadukova et al., 2008, Selinus, 2013, Kim et al., 2015).
Metal bioavailability is almost always less than the total metal concentration in the soil (Kirkham, 2006, Fernández-Calviño et al., 2017) and usually between 1 and 49% of total soil metal concentration may be available to plants depending on soil conditions (Naidu et al., 2003b, Selinus, 2013). It is influenced by physical, chemical and biological factors (Ernst, 1996, Harmsen, 2007, Gautam et al., 2017) including pH, soil type, redox potential, moisture status, metal species, the presence of oxides of Al, Fe and Mn, and the concentration of organic matter (Jjemba, 2005).
2.1.2.1 pH The pH of soil is a consequence of the H+ concentration in the solution present in soil pores, which is in dynamic equilibrium with the predominantly negatively-charged surfaces of the soil particles (Alloway, 1990b, Tukura et al., 2007, Bhat et al., 2011, Selinus, 2013). pH is seen as a master driver that influences all the factors determining metal bioavailability including the surface charge of silicate and other mineral clay particles, organic matter and oxides of Fe and Al, the sorption of metals and complexation with organic matter, the precipitation-dissolution reactions, redox reactions, ion mobility and leaching, and the dispersion of colloids (Adriano, 2001, Rengel, 2015). In general, the retention capacity of soils for metals increases with increasing pH up to about pH 7 (Figure 2-2). Exceptions are As, Mo, Se, V, and Cr, which are commonly more mobile (absorbed less completely) under alkaline conditions (Alloway, 1990b, Siegel, 2002, Selinus, 2013).
Figure 2-2 Modelled adsorption of certain trace metals onto hydrous ferric oxide (Source: Adriano, 2001). 39
For example, adsorption of Pb onto soils, such as by manganese oxides, increases as the pH increases (McKenzie, 1980, Martinez-Villegas et al., 2004) so that Pb mobility and bioavailability are reduced (Martinez-Villegas et al., 2004). Likewise, the mobility of Al increases significantly and it competes very actively with other cations for exchange sites in soil with pH below 5.5. In acid soils, the mobile Al can be absorbed quickly by plants and result in restricted root growth, drought susceptibility and inefficient metabolism of soil nutrients (Kabata-Pendias, 2001, Imadi et al., 2015, Kushwaha et al., 2016).
2.1.2.2 Cation Exchange Capacity (CEC) Cation exchange capacity is the number of moles of cations that can be adsorbed by a defined quantity of soil and it influences the distribution of positively charged ions between the soil particles and the soil solution. Most metals (with certain exceptions, including the metalloids As, Sb and Se and the metals Mo and V) exist primarily as cations in the soil solution. Therefore, their adsorption depends on the density of negative charges on the surfaces of the soil colloids (Alloway, 1990b, Siegel, 2002, Selinus, 2013). The CEC of soils is mainly dependent on the quantity and type of clay, organic matter, and the oxides of Fe, Al, and Mn, which have different cation exchange properties. In general, the higher the CEC of a soil, the greater the quantity of metals a soil can retain without potential hazards (Adriano, 2001, Selinus, 2013).
2.1.2.3 Soil Clay Minerals and Iron, Aluminium and Manganese Oxides The contribution of clay minerals to soil chemical properties results from their comparatively large surface area, permanent surface negative charge, and consequently high CEC (Alloway, 1990b, Caravaca et al., 1999, Huang and Matzner, 2007, Abollino et al., 2008, Selinus, 2013), so that increasing clay and hydrous oxide (Fe, Al and Mn oxides) contents in soils provide more sites for adsorption of metals and thus reduce metal bioavailability (McKenzie, 1980). For example, with increasing clay content As retention increases (Dickens and Hiltbold, 1967) and mobility and bioavailability decrease (Adriano, 2001). Iron, Mn and Al hydrous oxides, which are found in considerable amounts in acid soils, submerged soils, and their clay fractions (Ghanem and Mikkelsen, 1988), are the primary soil components responsible for As sorption (Huang and Matzner, 2007). In general, the sorption of As(V) is greater for both Fe and Al hydroxides than is the sorption of As(III) (O'Neill, 1990). McKenzie (1980) observed that adsorption of Pb2+ by Mn oxides was up to 40 times greater than that by the Fe oxides.
2.1.2.4 Soil Organic Matter Metal ions can be complexed by organic matter, changing their availability to plants. The COO - groups in both solid and dissolved organic matter form stable complexes with metals (Stevenson, 40
1976, Fellman et al., 2008, ElBishlawi and Jaffe, 2015, Reichman, 2002). Thus, increased soil organic matter content increases the opportunity for forming stable metal-organic matter complexes. In general, plants are unable to absorb the large metal-organic complexes and so the bioavailability of metals decreases (Moreno-Jiménez et al., 2017, Reichman, 2002). Because of the strong affinity of Pb for organic matter and its generally immobile nature, total Pb concentrations are often significantly higher in the surface than in subsurface soil horizons, due principally to plant dry matter recycling (Adriano, 2001). Similarly, Al toxicity may be reduced in the presence of organic matter (Vieira et al., 2009, Qin et al., 2011, Siecińska and Nosalewicz, 2017). In contrast, high organic matter content in soils promotes both As(III) and As(V) solubilisation in soils (Dobran and Zagury, 2006) as organic matter (fulvic or humic acids) forms stable complexes with mineral surfaces, effectively blocking arsenic from adsorption on iron oxides, alumina, quartz or kaolinite (Bauer and Blodau, 2006). Decomposition of organic matter, particularly that enriched in ammonium (e.g. sewage sludge), increases metal solubility and bioavailability and may induce Mn phytotoxicity (Adriano, 2001, Ingelmo et al., 2012).
2.1.2.5 Redox Potential Soils are subject to variations in oxidation-reduction (redox) status due to many factors including waterlogging and compaction (Alloway, 1990b, Herbauts et al., 1996, Engelaar et al., 2000, Selinus, 2013). Because metals are mainly adsorbed onto Fe-Mn oxyhydroxides (Chuan et al., 1996, Pan et al., 2016), redox potential can influence the availability and adsorption in soils by affecting the nature and number of active metal-binding sites in soil (Violante et al., 2010, Rengel, 2015). Under reducing conditions, Mn-oxides and Fe-oxides solubilize decreasing potential metal adsorption sites (Charlatchka and Cambier, 2000, Miao et al., 2006). Soil redox conditions also affect the proportions of particular metal species (e.g. Mn(II) versus Mn(IV), or As(III) versus As(V)) and the solubility of metals in the soil solution (e.g. Pb, Cd and Zn; (Chuan et al., 1996)). Generally, the solubility of metals is highest in acidic and reducing conditions, but pH has a stronger effect on metal solubility than does redox potential (Rengel, 2015). Increased As solubility under reducing conditions is associated with dissolution of Fe and Mn oxides/hydroxides (Carbonell-Barrachina et al., 2000, Fitz and Wenzel, 2002). In oxidizing conditions As usually exists as arsenates As(V)
(salts of orthoarsenic acid H3AsO4), whereas in mildly reducing conditions, As is generally present as arsenites As(III) (salts of arsenious acid H3AsO3) (Jones, 2007). In moderate reducing conditions, As often combines with S and Fe to form AsS or FeAsS, which are virtually insoluble in water and immobilized in the environment. In strongly reducing environments, elemental As (0) or H3As (-3) can exist, but such conditions are rare (Carbonell-Barrachina et al., 2000, Jones, 2007). As(III) predominates in reducing conditions and has been reported to be 4 to 10 times more soluble in soils
41 than As(V). In rice, translocation of Pb from roots to the shoot increases with increasing redox potential (Adriano, 2001). Mn(III) and Mn(IV) oxidation states occur as precipitates in oxidised alkaline environments, whereas the Mn(II) oxidation state is dominant in solution and solid phases under reducing conditions (Adriano, 2001).
2.1.2.6 Aging Aging of elements in soils has a significant influence on mobility and bioavailability of metals to plants (Lu et al., 2009, Huang et al., 2015). In soils, metals undergo organic complexation and, like most organic pollutants, are subject to aging processes that tend to immobilize them and reduce their bioavailability (Jjemba, 2005, Selinus, 2013). With time, elements added to soils are slowly relocated into less-accessible soil compartments, from which they are less readily exchanged and mobilized back into solution (Antoniadis et al., 2017). These elements also progressively form inner-sphere surface complexes with clay lattices (Antoniadis et al., 2017). Also, after the addition of metal to a soil, its availability decreases with time, usually because of micropore diffusion, cavity entrapment, occlusion in solid phases by co-presipitation and co-flocculation, surface precipitation or crystal growth (Ma et al., 2006a). Egodawatta et al. (2018) claimed that bioavailability of As and Sb depends on soil properties, total soil metalloid concentrations and incubation time, and that percentage of As and Sb bioavailability in historically contaminated soils (~34 years) was <10% and <2%, respectively, and varying little over wide ranges of total soil elemental concentrations, compared to <40% and <65%, respectively, in recently contaminated soils (aged for 14 days), even at low total soil elemental concentrations. Liang et al. (2014) reported that, during aging period, As transformed from non-specifically and specifically sorbed fractions (more available forms) to amorphous and crystallized Fe/Al fractions (less available forms) while Pb changed from exchangeable, carbonate and Fe/Mn hydroxide to organic fractions. Zn cations added to soils are absorbed quickly on the soil surfaces and then gradually diffuse into micropores, from where their availability is reduced (Ma and Uren, 2006). Sayen and Guillon (2014) reported that freshly added Zn was mostly present in an exchangeable fraction but it decreased over time, with 45% of Zn being remobilized into stronger binding sites after 63 days of aging. As a result, in experiments in which metals have been recently added to soils, the availability of metal is greater than in those in which the metals have aged for some period (Sauve, 2002, McBride and Cai, 2016, Egodawatta et al., 2018).
Aging also had significant effects on the concentrations of toxic elements in plants. For example, in soils with recently added elements, Cherif et al. (2011) observed that leaves of tomato plants irrigated with Cd solutions of up 1.12 mg Cd L−1 contained >30 mg Cd kg−1 while Amer et al. (2013) reported that addition of 100 mg L−1 of Pb and Zn resulted in >600 mg Pb kg−1 and >4200 42 mg Zn kg−1 in aerial parts of Medicago lupulina and Atriplex halimus, respectively. In contrast, various studies have shown that plants grown on soils where elements have been present or dumped for decades, contained low element concentrations (Chen et al., 2014, Ferri et al., 2015, Liu et al., 2016).
There are several factors affecting the rate of element aging. These include pH, organic matter content, incubation time, temperature (Ma et al., 2006b) and clay mineral content (Kim et al., 2008, McBride and Cai, 2016). Aging of elements was shown to be primarily governed by soil pH, possibly because of precipitation and diffusion (Ma et al., 2006b, Zhang et al., 2017a). Metal solubility is likely to increase at lower pH and decrease at higher pH conditions (Rieuwerts et al., 1998). Ma et al. (2006b) observed that attenuation rates of added Cu in soils treated with 2-48 g Cu L-1 (to reduce plant growth by 90 per cent) were higher compared to soils treated with 0.2-12 g Cu -1 L (to reduce plant growth by 10 per cent), especially in the initial phase, 1 and 5 d after metal addition. This result was because of the greater ease of precipitation and coagulation (occlusion) of colloids in the soils with higher added Cu concentrations (Ma et al., 2006b). This implied the lability of Cu added to soils declined quickly after the addition, particularly in soils with pH >6.0 followed by a gradual decrease in Cu lability (Ma et al., 2006b). Clay mineral content also had a significant role in governing the availability of added metals in soils over time as a greater aging effect on Cu toxicity was detected for fine-textured than coarse-textured soils (Kim et al., 2008, McBride and Cai, 2016).
2.1.2.7 Type and Speciation of Metals Metals are present in a number of forms in soils. Some forms or species are highly soluble, while others are so inert that their presence hardly influences the amount of the metal that is present in the soil solution phase (Sauve, 2002, Selinus, 2013). The distribution of these forms or species and their partitioning between the soil matrix and solution are not constant, but vary according to all the factors discussed above (Sauve, 2002, Krami et al., 2013).
The form of a metal in the soil can affect its phytoavailability and phytoxicity, especially for arsenic (Marin et al., 1992, Moreno-Jiménez et al., 2012). For example, As mostly occurs in soils as arsenate As(V) and arsenite As(III), which are considered to be the most toxic forms to plants (Adriano, 2001, Tripathi et al., 2007). Phytotoxicity of As(III) and As(V) depends on its oxidation state where As(V) is the predominant form in soils under oxidizing conditions while As(III) is the most stable and primary form under reducing conditions (Ruokolainen et al., 2000, Dobran et al., 2006, Anawar et al., 2018). Besides redox potential, pH and soil microorganisms also have a role in the interconversion of As(III) and As(V) (Zhao et al., 2010, Nearing et al., 2014, Farooq et al., 43
2016). Concentrations and solubility of As in soils are also mainly controlled by arsenic sorptivity by soil components, such as soil Fe, Al, and clay content (Walsh et al., 1977, Smith et al., 1999, Sahu et al., 2012). Both As(III) and As(V) are readily taken up by cells of the plant root (Finnegan and Chen, 2012). These As forms disrupt plant metabolism through different mechanisms where As(III) binds to and potentially deactivates enzymes comprising closely spaced cysteine residues while As(V) is a phosphate analog and can interrupt some phosphate dependent aspects of metabolism (Finnegan and Chen, 2012). Plant tissue analysis of root and shoot showed that 90% of As was detected in form of As(III) even though plants were treated with As(V) (Farooq et al., 2016). This could occur because of transformation of As(V) to As(III) after it reaches the cell (Daus 2008, Finnegan and Chen, 2012). Both As(III) and As(V) reduce seed germination, root and shoot length (Marschner et al., 2002, Ahmad and Gupta, 2013, Sanal et al., 2014, Yoon et al., 2015), inhibite plant growth, disrupt photosynthetic and respiratory systems (Garg and Singla, 2011, Dehabadi et al., 2012), and decrease chlorophyll and protein content (Ahmad and Gupta, 2013). However, As(III) is generally regarded as the most soluble and mobile in the environment as well as the most toxic species to plants, such as Tristicum aestivum (Liu et al., 2005), Brassica juncea (Pickering et al., 2000, Ahmad and Gupta, 2013), and Holcus lanatus (Quaghebeur and Rengel, 2003), followed by As(V) (Liu et al., 2005, Jones, 2007). As shown in Figure 2-3, at the same arsenic concentration (16 mg.L-1), arsenite reduced the mean germination percentage of six wheat varieties by 38% compared to 24.2% for arsenate. Chandra et al. (2016) also observed that at 50 mg As.kg-1, arsenite reduced shoot height, root length, shoot biomass and root biomass of Abelmoschus esculentus by 55.16%, 61.9%, 63.77% and 83.27%, respectively, relatively to the control, while the reduction was by 26.01%, 42.86%, 47.07% and 65.98%, respectively, in arsenate treatment. Various studies have also reported more detrimental effect of As(III) on many morphological, physiological, and biochemical changes of numerous plants compared to As(V) (Srivastava et al., 2007, Ahmad and Gupta, 2013, Sanal et al., 2014, Yoon et al., 2015, Armendariz et al., 2016). Daus (2008) found that As(III) is more toxic to animals and humans than As(V) but the reverse was found for plants which can reduce arsenate to arsenite (Daus, 2008, Finnegan and Chen, 2012, Meadows, 2014). Vromman et al. (2018) demonstrated that As(III) was more toxic than As(V) to Atriplex atacamensis plant at the plant level as it reduced plant growth, stomatal conductance and photosystem II efficiency but it was not more toxic than As(V) at cellular level. Gusman et al. (2013) observed both As(III) and As(V) had the same toxic effects in leaves and roots of Lactuca sativa seedlings. Abbas and Meharg (2008) observed more toxic effect of As(V) on Zea mays compared to As(III). Plants differ in their capacity to immobilize As as different species may have different metabolic tolerance to As (Yoon et al., 2015, Finnegan and Chen, 2012). Besides metalloid speciation, toxicity of As to plants differs depending on plant species or varieties and soil 44 factors governing As accumulation in plants (Abbas et al., 2018, Yoon et al., 2015). In addition, different species of plants have different mechanisms of As(III) or As(V) uptake, toxic response, and detoxification (Abbas et al., 2018). Therefore, understanding of these mechanisms of transformation between the two forms of As (in soil and in plants) is crucial to predicting the extent of, and for dealing with their toxicity (Zhao et al., 2010, Flora, 2015). The chemistry of Cr is also relevant to biology, ecology and human health. Of the two forms found in nature, Cr(VI) is more toxic than Cr(III) (Adriano, 2001). Aluminium may be present in several chemical forms, with different toxicities. The trivalent Al species (Al3+) dominates in acid soils and is the major phytotoxic species (Delhaize and Ryan, 1995, Ryan and Delhaize, 2010).
Figure 2-3 Average germination percentage of six varieties of wheat growing in different concentrations of arsenite () and arsenate () (Source: Liu et al., 2005).
2.1.2.8 Soil Amendments Numerous contaminated soil amendments have been studied to determine their efficiency for the immobilisation of metals. These included natural zeolites (Erdem et al., 2004), fly ash (Gupta et al., 2000, Bertocchi et al., 2006), bauxite red mud (Brunori et al., 2005, Bertocchi et al., 2006), sewerage sludge, and animal manure (Wong and Lau, 1985, Schwab et al., 2007). Organic amendments such as manure were found to be effective in reducing lead availability in mine tailings (Scialdone et al., 1980, Wong and Lau, 1985, Ye et al., 1999). They also offer a slow release of nutrients, promote soil structural stability, and increase the CEC and organic carbon content of the soil (Clemente and Bernal, 2006). As a result, organic amendments may promote establishment of vegetation in mining areas which then reduces metal contamination via runoff and erosion (Schwab et al., 2007, Barajas-Aceves and Rodríguez-Vázquez, 2013, Yuan et al., 2016). Nevertheless, organic amendments can increase the risk of metal leaching. For example, Schwab et al. (2007)
45 found that the maximum concentrations of Zn in leachate from manure-amended treatments were up to 3.7 mg.L-1, whereas control Zn concentrations were less than 0.7 mg.L-1. The presence of aged cattle manure significantly increased Pb concentration in soil leachate from less than 10 µg.L-1 to higher than 60 µg.L-1 (Schwab et al., 2007).
Lime has been used to increase the soil pH to close to neutral and thereby reduce metal mobility (García-Sánchez et al., 1999, Hale et al., 2012). It is very effective in immobilizing Pb in contaminated soils through precipitation of Pb hydroxides (88% immobilization of Pb at pH 12.6 using a mixing ratio of lime: soil of 1:21) (Alpaslan and Yukselen, 2002). Agricultural lime, in which the principal cation is calcium, can also immobilize As in contaminated soils and sludges by the formation of calcium arsenate (Ca-As) precipitates (Vandecasteele et al., 2002, Moon et al., 2004). However, Jones et al. (1997) stated that lime could also mobilize As due to the pH dependence of the As sorption reaction on oxide minerals and layered silicates (clays). Codling and Ritchie (2005) observed a slight and insignificant increase in tissue As concentration but a reduction in tissue Pb concentration of eastern gamagrass (Tripsacum dactyloides L.) grown on As and Pb contaminated soils that were amended with lime and phosphorus. Unlike Pb, the effect of lime on immobilization of As is inconsistent.
A wide range of polymers has been used to ameliorate arid conditions, as well as saline- or metal- contaminated soils (Rossato et al., 2009, Rossato et al., 2011, Park et al., 2012, Shahid et al., 2012). For instance, Guiwei et al. (2008) investigated the effects of different application rates of insoluble hydrophilic polyacrylate polymers (0.2; 0.4; and 0.6% amendment) on plant growth and soil quality of a Pb-contaminated mine soil. They found that the polymer increased the water holding capacity of the soil by almost four times and decreased bioavailable Pb by 34 to 85% as compared to the unamended soil. In addition, the biomass of the orchardgrass (Dactylis glomerata L. cv. Amba) plants grown in soil amended with 0.4% polymer was more than 3000 times higher than control plants grown on unamended soil. However, the capacity of the polymer to retain water decreased progressively by about one-quarter over a 100-day experiment, presumably as the polymer sorbed Pb (Guiwei et al., 2008). Banedjschafie and Durner (2015) also revealed that the water holding capacity of cross-linked polymers in sand deceased by about half during a period of six months, which they attributed to the accumulation of salts. In addition, soil enzymatic activities such as urease (which catalyses the hydrolysis of urea to CO2 and NH3) were inhibited by the application of polymer (de Varennes et al., 2010) suggesting potential toxicity.
46
2.1.3 Determination of metals in soils (and plants) To assess the toxicity of metals to plants, analytical approaches must be able to quantify the fraction of the total metals in a soil which is bioavailable, i.e. the fraction that is available for uptake by plants or soil organisms. As discussed in Section 2.1.2., around 1 to 49% of total metals are phytoavailable depending on the total metal concentration and soil pH.
Rao et al. (2008) indicated five principal metal extraction methods (Figure 2-4): Column leaching with artificial rainwater. Single extraction of a soil sample using unbuffered salt solutions, such as calcium chloride
(CaCl2), sodium nitrate (NaNO3) or ammonium nitrate (NH4NO3). Sequential extraction of a sample of a soil or sediment with a series of reagents in order to partition trace element compounds of differing solubilities. Pseudo-total soil content of samples digested using strong acid or aqua regia. This method indicates the maximum potentially soluble or mobile contents of metals and, in the case of environmental metal contaminants, those usually not bound in silicates. It is a measure of the maximum potential hazard that could occur in the long term or in extreme environmental regimes. Total metal content of samples digested using acid such as hydrofluoric acid (HF) and
perchloric acid (HClO4) to break down all compounds, including silicates.
47
Figure 2-4 Different approaches applied in general to the determination of metals in soils (% values shown are approximate as they are metal and matrix dependent) (Source: Rao et al., 2008).
Sequential extraction of minerals from soil quantifies the proportions held by a range of physical and chemical phenomena (Rao et al., 2008, Zimmerman and Weindorf, 2010, Kim et al., 2015). Table 2-1 presents generalised categories of metal occurrence, extraction methods and the incremental percentage of metal released by the different methods if used sequentially.
48
Table 2-1 Categories of metal occurrence in soil, typical sequential extraction methods and typical percentage of total metal released (Compiled from Rao et al., 2008, Zimmerman and Weindorf, 2010, Kim et al., 2015). Typical % Metal occurrence category Typical extraction method total metal a. In soil solution: as free metal ions and Column leaching with artificial 5 soluble metal complexes rainwater b1. Adsorbed to inorganic soil constituents Unbuffered salt solutions: CaCl2, 5
at ion exchange sites NaNO3, NH4NO3 pH 7 b2. Bound to oxides, carbonates Complexing agent: EDTA, DTPA 10 c1. Bound to soil organic matter H2O2, HNO3, NH4COOH pH 2 20 c2. Precipitated as oxides, hydroxides, and HCOOH, NaCOOH pH 2 20 carbonates d. Residual HNO3+HCl 30 e. Embedded in silicate mineral crystal HF+HClO4 10 structure
2.1.3.1 Methods for measuring total metal in soils and plants Methods used in Australian laboratories for determining total plant and soil metal contents were reviewed by Handson and Shelley (1993) and by Rayment and Lyons (2011). Nitric-perchloric, nitric-sulfuric, and nitric-perchloric-sulfuric acid mixtures are the most frequently used extraction methods. In some cases hydrogen peroxide is also used to reduce the occurrence of undigested black particles found with some samples such as Coal Fly Ash (Wu et al., 1996) and also to reduce the carbon content in samples and to maintain a higher temperature at a safe working pressure (Wu et al., 1997). Other extraction methods include mixtures of nitric acid, hydrochloric acid and hydrofluoric acid. Microwave digestion is now routinely used in soil and plant testing laboratories (Davidson, 2013).
The combined use of HF, HClO4 and HNO3 appears to be the most effective method for determining total metals in soils and plants as samples are digested completely (Rao et al., 2008, Davidson, 2013). Dissolution with HF is important because it attacks silica and silicates which form a large part of the soil matrix (Ure, 1990) and are also present in plants, particularly grasses (Hunt et al., 2008), and because metals can be bound strongly to them.
49
2.1.3.2 Measures of bioavailable metals in soils Regardless of its environmental importance, there is no consensus in the literature as to which extractants most accurately estimate the phytoavailability of trace metals in soils (Menzies et al., 2007, Soriano-Disla et al., 2010, Zhang et al., 2010, Kim et al., 2015). There is little consistency between researchers in the selection of single extractants for bioavailable metals. All workers assume that soluble metals ions (Category a) (Table 2-1) are bioavailable, but there are differences between the ions that are removed by neutral inorganic salts (Category b1) (Table 2-1) and those removed by organic complexing agents (Category b2) (Table 2-1). Metal ions released by acid treatment from organic or inorganic compounds (Category c) (Table 2-1) are sometimes included in the bioavailable fraction of total metal concentration. There are important practical reasons for using a single extractant, particularly if the final analyses are to be undertaken by a method such as inductively coupled plasma spectroscopy. Several of the more common single extractants that have been used to determine availability of metals to plants are shown in Table 2-2.
50
Table 2-2 Examples of extractants used for assessing plant-available trace elements concentration in soil (from numbered sources, 1-27 cited at the bottom of the Table). Extractant Element Ammonium Ammonium Sodium Calcium Water DTPA EDTA acetate nitrate nitrate chloride Al 5 3, 5 5, 7 5, 19, 26
As 8 18 8, 15, 17, 20 20 15
1, 2, 6, 12, 13, 1, 2, 1, 2, Cd 4, 5 5, 13 5, 7, 16, 18 11, 16 16, 17, 19 16 13, 16 Co 24 16, 25 16 16, 24, 25 16, 24 16 Cr 13 7, 16 16 13, 16, 19 16 13, 16
1, 3, 5, 12, 13, 1, 2, 1, 2, Cu 4, 5 5, 6, 13 5, 7, 16 11, 16 14, 16, 17, 19, 3, 16 13, 16 27 2, 3, Fe 3, 6, 13 16 16 3, 13, 16, 19 13, 16 16 Hg 21 7 21, 22, 23 21, 23 23
La 5 5 5, 7 5
3, 10, 13, 16, 17, 2, 6, 2, 9, Mn 6, 9, 13 7, 9, 16 16 19, 26 16 13, 16 Mo 7 3, 17
Ni 13 7, 12, 16 11, 16 1, 13, 16, 17, 19 1, 16 13, 16
1, 10, 13, 16, 17, 2, 16, 9, 13, Pb 9, 13 7, 9, 16, 18 11, 16 19, 20, 26 20 16 Sb 20 20 Y 5 5 5, 7 5
1, 3, 5, 6, 12, 13, 1, 2, 2, 13, Zn 4, 5 5, 6, 13 5, 7, 16, 18 11, 16 14, 16, 17, 19, 3, 16 16 26, 27 Reference abbreviations: 1, Menzies et al. (2007); 2, Smilde et al. (1992); 3, Rayment and Lyons (2011); 4, Mahler et al. (1980); 5, Takeda et al. (2006); 6, Sterckeman et al. (1996); 7, Rékási and Filep (2006); 8, Száková et al. (2005); 9, Grønflaten and Steinnes (2005); 10, Gupta and Sinha (2006a); 11, Gupta and Aten (1993); 12, Merkel (1996); 13, Sakan et al. (2016); 14, Luo and Rimmer (1995); 15, Tan et al. (2018); 16, Gupta and Sinha (2007); 17, McBride et al. (2003); 18, Neu et al. (2018); 19 Wightwick et al. (2010); 20, Wanat et al. (2014); 21, Jin et al. (2008); 22, Ma et al. (2015); 23, Rodríguez et al. (2017); 24, Hsiao et al. (2009); 25, Wendling et al. (2009); 26, Anjos et al. (2012); 27, McBride and Cai (2016).
The different extractants used to determine bioavailability of metals in soils operate most effectively under a wide range of conditions, but neutral salt extractants such as CaCl2 and NaNO3 might provide the best measure of metal phytoavailability across a large range of metals at near- neutral pH, including Pb (Pueyo et al., 2004, Gupta and Sinha, 2006b, Menzies et al., 2007), As 51
(Száková et al., 2005, Rao et al., 2008), and Al and Cu (Takeda et al., 2006), Cd, Ni and Zn (Kim et al., 2015). They are suited for the prediction of plant metal uptake in soils and can be considered as good models for simulating raining and flooding events (Rao et al., 2008). They may also provide the best relationship between soil extractable trace metals and plant tissue accumulation (Rao et al.,
2008). Of these salts, 0.01 M CaCl2 the most widely used extractant (Table 2-2) is recommended for general use (Menzies et al., 2007, Kim et al., 2015).
2.2 Metals in plants 2.2.1 Metal uptake by plants There are two mechanisms of metal transport from the bulk-soil to plant roots: 1) mass flow or convection, and 2) diffusion (Lasat, 2000). Convection carries soluble metal ions from the soil matrix to the root surface with water that replaces transpirational losses from leaves. Some ions are absorbed by roots faster than the rate of supply through mass flow and this creates a depleted zone in the soil around the root and generates a concentration gradient which drives the diffusion of ions toward the roots (Clarkson, 1993a).
Metal uptake at the root surface combines diffusion and physiological uptake processes within root cells (Marschner, 2012). Metals may also enter the plant roots via uptake of metal by symbiotic microorganisms associated with the root and transfer of metal from the symbiont to the root (Lasat, 2002, McLaughlin, 2002, Audet and Charest, 2009). Some microorganisms may excrete organic compounds which increase bioavailability and facilitate root absorption of metals such as Fe, Mn and Cd. Soil microorganisms can also directly influence metal solubility by changing their chemical properties, for example, a strain of Pseudomonas maltophilia was shown to minimize environmental mobility of toxic ions such as Hg, Pb and Cd (Lasat, 2000).
Ions move from the root epidermis to the xylem vessels by apoplastic or symplasmic flow (McLaughlin, 2002). Apoplastic flow occurs when ions move through the space within and between the cell walls. Symplasmic flow occurs when metals cross the plasma membrane to the cytoplasm of cells in the cortex and move via the plasmodesmata through the Casparian strip to the stele (Figure 2-5).
Metal ion transport from the apoplast to the symplasm must be mediated by membrane proteins with transport functions, known as transporters, as the ions cannot move freely across the lipophilic cellular membranes (Clarkson, 1993a, Lasat, 2000, Manara, 2012). These transporters are characterized by kinetic parameters, such as maximum uptake capacity (Vmax) and affinity for a metal ion (Km). For example, arsenate (V) enters plant roots on a phosphate-transport protein as it is
52 a P analog (Meharg and Macnair, 1992a, Poynton et al., 2004, Danh et al., 2014). Arsenate (As(V)) influx can be described by Michaelis-Menten kinetics and the kinetic parameter Km was found to be lower in the As hyperaccumulating species Chinese brake fern (Pteris vittata L.) and Cretan brake (Pteris cretica cv. Mayii L.) than in a non-accumulating fern (Nephrolepis exaltata L.), indicating higher affinity of the transport protein for arsenate. Phosphate inhibits arsenate influx in a directly competitive manner (Poynton et al., 2004). Wang et al. (2002) observed that the removal of As(V) from solution by P. vittata was inhibited by increasing the supply of P. It was also shown that phosphate alleviated arsenate toxicity in non-tolerant common velvet grass (Holcus lanatus) over extended periods of time (Meharg and Macnair, 1992a). This is consistent with the hypothesis that arsenate enters plant roots on a phosphate-transport protein. In contrast, Christopherson et al. (2009) did not detect direct competition between the uptake of P and As(V) by medic (Medicago truncatula) and barley (Hordeum vulgare) grown in soils rather than solution culture, even though the addition of P ameliorated the effects of As(V) on plant growth. This indicates a complex interaction between elemental concentrations, uptake patterns and internal effects within the plant.
Figure 2-5 Simplified structure of the root, showing apoplastic (left) and symplastic (right) pathways of radial metal transport and exchanges between apoplast and symplast (bottom) between the root surface and the stele (Source: McLaughlin, 2002).
2.2.2 Metal translocation to shoots
Only a part of the ions associated with the root is absorbed into cells. A significant ion fraction is adsorbed at the extracellular negatively charged sites (COO-) of the root cell walls which are responsible for immobilization of metals in roots and subsequent inhibition of ion translocation to the shoot (Clarkson, 1988, Krämer, 2010). For example, a wide range of plants accumulate Pb in roots, but Pb translocation to shoots is very limited (Kopittke et al., 2007).
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The degree and extent of ion movement within plants are based on the metal concerned, the plant part and plant age (Furini, 2012). Elements such as Mn, Zn, Cd, B, Mo and Se are readily translocated to the plant tops; Cu, Ni and Co are intermediate; Pb, Cr, Hg are transferred to the least extent (Zitka et al., 2013). Concentrations of As in roots of terrestrial plants are normally higher than in the stem, leaves or fruit (Alloway, 1990b, Adriano, 2001, Jedynak et al., 2009, Selinus, 2013) although Chinese brake fern (Pteris vittata L.) (an As hyperaccumulator) was found to contain more than 6000 mg.kg-1 of As in its frond dry mass (Tu et al., 2002). Aluminium is not readily translocated from root to shoot (Ma and Hiradate, 2000). However, some plant species are able to accumulate Al to high concentrations in shoots with no toxicity symptoms, such as tea (Camellia sinensis L.) which accumulated Al up to 30,000 mg.kg-1 DW in its older leaves (Matsumoto et al., 1976).
2.2.3 Symptoms of metal toxicity in plants Some metals (B, Co, Cu, Fe, Mn, Mo, Ni and Zn) are essential for plant growth in small amounts (micronutrients) (Furini, 2012, Gupta et al., 2013). Metals, such as As and Pb, are not essential for plant growth and are toxic to plants even at low concentrations (Furini, 2012).
Plant growth responds directly to increasing concentrations of most metals; they may show deficiency responses when essential metal concentrations are low, no effect when it is sufficient and decline in growth when the metal concentrations are higher (toxic) (Marschner, 2012). Toxicity threshold concentrations and toxicity symptoms vary with the metal and also with the plant species (Jones, 1998, Selinus, 2013). Non-essential heavy metals such as Cd, Hg, Ag, Pb and U are toxic even at very low concentrations (Furini, 2012, Gupta et al., 2013).
All metals, essential or non-essential, are toxic to plants at high concentrations, at which plants show symptoms of metal toxicity (Mallick and Rai, 2002). The critical metal concentration in plant tissues or the soil solution is the metal concentration necessary for toxicity symptoms to become apparent, for example just before plant dry matter is reduced (Baker and Walker, 1990), or when there is a 10% reduction in growth or yield (Davis et al., 1978, Aery et al., 1997, Zhang et al., 1998, Reichman et al., 2002). The critical concentrations may be extremely low for highly toxic metals such as Pb and As. Effects of toxic metal concentrations (e.g. Cu) in soils may be reduced if plants are grown at high density, due to a reduction in the effective dose to which each plant is exposed (Hansi et al., 2014). The metal toxicity limits in plants and in soil for several metals are described in Table 2-3.
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Table 2-3 Metal toxicity limits. Toxicity index Soil solution plant Soil plant toxicity Plant leaf tissue Metals toxicity concentrationa toxicity limitsc concentrationb (mg.kg-1) (mg.kg-1) (mg.L-1) Al - 100 200 As 15 0.67 5 – 20 Cd 3 0.56 5 – 30 Co - 1 2 Cu 200 0.13 2 – 20 Hg - 0.09 0.3 Mn 3,000 2.53 400 – 1,000 Ni 9 1.12 10 – 100 Pb 100 – 500 0.06 30 – 100 Zn 400 1.63 100 – 400 aBased on total metal concentrations generally toxic to plant growth; bBased on metal concentrations in soil solution generally toxic to plant growth. cBased on mean values of toxic concentration of metals accumulated in agricultural crops (Sources: Mendez and Maier, 2008a, Kopittke et al., 2010).
2.2.3.1 Effect on seed germination and early root growth The seed is a phase in the plant life cycle which is highly protected against a wide range of external stresses. In contrast, imbibition and the subsequent processes of germination and early vegetative development are commonly very sensitive to stress including the presence of metals as resistance mechanisms are less developed in germinating seeds than in mature plants (Liu et al., 2005, Kranner and Colville, 2011).
The most common symptoms of metal toxicity in germinating seeds are reductions in germination percentage, shoot and root length (Xiong, 1998, Kranner and Colville, 2011). For example, the germination percentage of rice (Oriza sativa L.) treated with 8 mg.L-1 of arsenate and arsenite was decreased by 61 and 83%, respectively (Abedin and Meharg, 2002). The germination percentage of wheat (Triticum aestivum L.) grown on 16 mg.L-1 of arsenate and arsenite were also reduced by 16.3 and 38%, respectively (Liu et al., 2005). Root growth of curly Mitchell grass (Astrebla
55 lappacea) grown on 667.36 μM of arsenate was totally inhibited compared to the control grown on deionised water (Figure 2-6).
Figure 2-6 Effect of increasing concentrations of arsenate on early root growth of A. lappacea (Source: Guterres et al., 2019).
Xiong (1998) observed the effect of Pb on the germination of Chinese cabbage (Brassica pekinensis Rupr.) and found that 125 µg.mL-1 of Pb significantly reduced the root length of the species by 50% whereas the germination rate was similar to that of the control. Barley (Hordeum vulgare L.) seedlings treated with Cd (1 mM), Ni (3 mM) or Hg (0.5 mM) showed root growth inhibition of approximately 45% as compared with the control (Tamas et al., 2008). Kaur et al. (2013) observed progressive reductions of up to 40% in root length of wheat (Triticum aestivum) seedlings as Pb concentrations increased from 0 to 500 μM but Lamhamdi et al. (2011) found no significant reduction in germination completeness or root length in wheat seedling roots at 300 μM Pb. The germination percentage and root growth of curly Mitchell grass (A. lappacea) treated with 9600 µM of Pb were reduced by 17.74 and 100%, respectively (Guterres et al., 2019).
Inhibition of root and shoot growth is also a visible symptom of Al toxicity. Al-induced inhibition of root elongation leads to root stunting (Mossor-Pietraszewska, 2001, Alvarez et al., 2012) and affected roots are unable to take up nutrients and water (Nosko et al., 1988). Silva et al. (2010)
56 investigated the uptake of Al and transport to shoots in Al-sensitive and tolerant wheat cultivars and found that Al exposure affected endodermis differentiation in the root hair region of the Al tolerant cultivar roots, resulting in thickening of all endodermal cells in the root hair zone after 24 h.
Meda and Furlani (2005) studied the tolerance of 17 species of tropical legumes to Al and concluded that root elongation was the best parameter to compare Al tolerance amongst the seedlings rather than shoot or root dry matter production.
On this basis, root length in germinated seedlings was selected as an indicator of metal tolerance during germination in our study.
2.2.3.2 Effects on plant growth The most common visual symptom of metal toxicity in plants is a reduction in plant growth as metal toxicity increases although the percentage of reduction differs among metals and plant species (Reichman, 2002, Peng et al., 2010, Selinus, 2013). Elevated concentration of arsenic (10 mg.L-1) reduced root plus shoot dry biomass production of tomato (Lycopersicum esculentum Mill) to 42.2% of the control and also caused death of bean (Phaseolus vulgaris L.) plants after 36 days (Carbonell-Barrachina et al., 1997). Gopal and Rizvi (2008) observed the effect of Pb on the growth of radish (Raphanus sativus) cv. Jaunpuri and found that excess Pb (0.5 mM) significantly reduced the dry weight of the species by 50% as compared to that of control 35 days after treatment began. Symptoms of Al toxicity in wheat (Triticum aestivum L.) were relatively short and thick roots with numerous undeveloped laterals (Taylor and Foy, 1985). Aluminium toxicity symptoms were also observed in tropical leguminous plants, for example darkening, shrinking and inhibition of lateral root growth two days after transplanting (Meda and Furlani, 2005). The critical mechanism appears to be the binding of Al to the pectic cell wall matrix and to the apoplastic face of the plasma membrane in the portion of the root apex that is particularly sensitive to Al, thereby impairing processes in both the symplast and apoplast that affect cell function and wall extension (Horst et al., 2010).
2.2.3.3 Toxicity symptoms in leaves Foliar symptoms of metal toxicity in plants are metal and species specific. At 10 µM, Cu caused restricted growth and reddening of leaf tips in narrowleaf red ironbark (Eucalyptus crebra), river red gum (Eucalyptus camaldulensis) and silver-leaf wattle (Acacia holosericea) (Figure 2-7) and bright red older leaves in weeping paperbark (Melaleuca leucadendra) (Reichman, 2001).
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(a) (b)
Figure 2-7 Copper toxicity (reddening of leaflet and stunted growth) in (a) Eucalyptus crebra (b) Acacia holosericea (Source: Reichman, 2001).
At 512 µM, Mn caused chlorosis and necrotic splotches in silver-leaf wattle (Acacia holosericea) and leaf curling and chlorotic or necrotic spotting in weeping paperbark (Melaleuca leucadendra). At 50 µM, Zn caused leaf bronzing, curling and tip necrosis in weeping paperbark (Melaleuca leucadendra) (Reichman, 2001). Leaf curling and wilting and red colouration of the old leaves were also observed in weeping paperbark (Melaleuca leucadendra) at 100 µM Zn (Figure 2-8).
(a) (b)
)
Figure 2-8 Zinc toxicity in Melaleuca leucadendra (a) Leaf curling and wilting (b) Red colouration of the old leaves (Source: Reichman, 2001).
Arsenic at 25 mg.kg-1 of soil can cause leaf chlorosis and necrosis around the edges of pinnae of tender brake fern (Pteris tremula), an As non-hyperaccumulator (Caille et al., 2005). Interveinal chlorosis along the margins of young leaves was observed in cabbage (Brassica oleracea L.) treated with 1 mM of Pb (Sinha et al., 2006).
Leaves of an Al-sensitive cultivar of maize showed purple colour and interveinal leaf chlorosis when treated with 111 µmol.L-1 of Al (Meda and Furlani, 2005). Similarly, younger leaves of coffee plants become small, curled along the margin and chlorotic while older leaves indicate a marginal 58 chlorosis which progresses to the centre of the leaf under Al stress conditions (Roy et al., 1988). The first specific foliar symptom of Al toxicity in wheat is a yellowing along the margin near the tip of the oldest leaf and within a few days brown lesions form in these chlorotic regions and extend in from the margins resulting in the formation of indentations (Figure 2-9) (International Maize and Wheat Improvement Centre, 2010).
Figure 2-9 Aluminium toxicity in Triticum aestivum L. (a) Stubby appearance and brownish colour of roots (b) Leaves showing yellowing/chlorosis along the margin and formation of necrotic indentations (Source: International Maize and Wheat Improvement Centre, 2010).
2.2.3.4 Other effects on plant physiology Toxic metals entering the plant tissues can inhibit most physiological processes at all levels of metabolism. These include reduction of photosynthesis, inhibition of enzyme and protein production and metabolism, interference of essential nutrient transport, including Cu, Zn, Ca, Mg, Mn, K, P and Fe, and negative effects on cellular functioning (cell division and cell elongation) and induction of DNA and cell death (Roy et al., 1988, Fodor, 2002, Tamas et al., 2008, Manara, 2012). Reduction of chlorophyll content can affect the whole plant metabolism through the functioning of the photosynthetic apparatus (Fodor, 2002). Singh and Ma (2006) observed metabolic adaptation to As-induced oxidative stress in Chinese brake fern (Pteris vittata L.) (As hyperaccumulator) and slender brake fern (Pteris ensiformis L.) (As non-hyperaccumulator) and found that at 133 µM As, concentrations of chlorophyll, protein and carotenoids increased in Pteris vittata L. whereas they drastically decreased in Pteris ensiformis L. In onion (Allium sepa), 10 mM Pb was shown to reduce root growth and to cause mitotic irregularities and chromosome stickiness (Liu et al., 1994). Zhang et al. (2007) studied the effect of Al in soil on photosynthesis and related morphological and physiological characteristics of two soybean genotypes and concluded that Al at 80 mg.kg-1 resulted in reduced chlorophyll contents, depressed photosynthesis and enhanced transpiration rates. A study on Al-sensitive and Al-tolerant maize also found that Al severely reduced photosystem activity and chlorophyll content of Al-sensitive maize (Mihailovic et al., 2008). In wheat, Al toxicity reduced 59 photosynthesis, chlorophyll concentration, and transpiration at critical Al concentrations in leaves of 0.15, 0.13, and 0.14 mmol.kg-1, respectively, whereas in sorghum, transpiration was increased by Al stress (Ohki, 1986).
2.3 Plant responses to metals Plants able to tolerate and grow on soil containing elevated metal concentrations are known as metallophytes. There are two categories of metallophytes according to their occurrence. Absolute metallophytes (strict or eumetallophytes) are plant species that are restricted to soils that contain relatively high metal concentrations or metal contaminated soils only. Pseudo (facultative) metallophytes occur on both high metal concentration or metal contaminated soils and on non- contaminated soils (Roosens et al., 2008, Epelde et al., 2010).
Responses of metallophyte plants to increasing concentrations of metals in soil can be classified into four categories, namely metal excluders, accumulators, hyperaccumulators and indicators (Baker, 1981), as described below and in Figure 2-10.
2.3.1 Metal excluders In this category, metal concentrations in the shoot are maintained constant and low over a wide range of total soil metal concentrations, up to a critical value above which the exclusion mechanism breaks down and unrestricted transport results (Baker, 1981). Excluders include some members of the grass family, Poaceae (e.g. sudangrass, bromegrass, fescue etc.), known for their insensitivity to metals over a wide range of soil concentrations and site conditions (Baker, 1981). Excluders survive metal toxicity through avoidance (or restriction) mechanisms (Baker, 1981).
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Figure 2-10 Relationships between plant tissue metal concentration and total metal concentrations in soil for the four different types of plant responses to metals (Source: Adriano, 2001).
2.3.2 Metal accumulators In metal accumulators, metals are concentrated in above-ground plant parts (above the threshold foliar metal dry weight concentrations for accumulators described in Table 2-4) from low or high soil concentrations (Baker et al., 1994a). Root uptake and transport of metals in these plants are more or less in balance but metals can accumulate in the roots (Baker, 1981). Members of the mustard (Brassicaceae, e.g. lettuce, spinach, chard, etc.), composite (Asteraceae) and tobacco (Solanaceae) families are often accumulator plants (Baker and Brooks, 1989) which are generally restricted to a particular type of soil and parent rock and may prove to be important geobotanical indicators for mineral deposits (Baker and Brooks, 1989). They survive metal toxicity through a tolerance mechanism and not through avoidance of metal toxicity (Baker and Brooks, 1989).
2.3.3 Metal hyperaccumulators Extreme metal accumulators are known as hyperaccumulators (Baker, 1981). In these plants, metal concentrations in shoots can be exceptionally high (about 100 times the highest values found in shoot of non-accumulating plants) (Singh and Ma, 2007) without any symptoms of physiological stress (Baker et al., 2000). Hyperaccumulators are classified by their threshold foliar metal dry weight concentrations which differ among metals (Table 2-4).
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Table 2-4 Threshold foliar metal dry weight concentrations in non-tolerant plants (agricultural crops and pastures) and metal accumulators and hyperaccumulators from literature. Normal range Accumulator Hyperaccumulator Metals mg.kg-1 mg.kg-1 mg.kg-1 % Zn 20 - 400 2000 10,000 1 Mn 20 - 400 2000 10,000 1 Fe 25 – 150 500 10,000 1 Ni 1 - 10 100 1,000 0.1 Cu 5 - 25 100 1,000 0.1 Co 0.03 - 2 20 1,000 0.1 Cr 0.2 - 5 50 1,000 0.1 Pb 0.1 - 5 100 1,000 0.1 As 0.009 – 1.5 40 1,000 0.1 Al 200 400 1,000 0.1 Cd 0.1 - 3 20 100 0.01 Se 0.05 - 1 10 100 0.01 Hg 0.01 – 0.3 3 10 0.001 Au 0.001 – 0.01 0.1 1 0.0001 (Kabata-Pendias, 2001, Kabata-Pendias and Mukherjee, 2007, Baker et al., 1994a, Brooks, 1987, Reeves, 1988, TotalGro, 2009, Chaudhry et al., 1998, Mossor-Pietraszewska, 2001, Anderson et al., 1998)
Hyperaccumulators are normally found in metalliferous soils (near ore deposits) or in heavily contaminated soils and survive metal toxicity through tolerance mechanisms (Baker, 1987, Adriano, 2001). They are low-biomass species as they use considerable energy in the processes of accumulation and tolerance of high concentrations of metals in their tissues (Kabata-Pendias, 2001). On a worldwide basis, more than 500 hyperaccumulator plant species have already been reported, most commonly from the Asteraceae, Brassicaceae, Caryophyllaceae, Cyperaceae, Cunoniaceae, Fabaceae, Flacourtiaceae, Lamiaceae, Poaceae, Violaceae, and Euphorbiaceae (Padmavathiamma and Li, 2007).
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2.3.4 Metal indicators Metal-indicator plant species respond proportionally to metal concentrations in soils, displaying linear relationships between tissue and soil metal concentrations (Baker, 1987). In these species, uptake and transport of metals to the shoot are balanced so that metal concentration in the shoot reflects metal concentration in the soil (Baker, 1987). Indicator species include the grain and cereals crops (corn, soybean, wheat, oats, etc.) (Baker, 1981).
2.4 Tolerance mechanisms in plants Plants growing on soils with elevated concentrations of metals possess a range of tolerance mechanisms that enable them to survive or avoid metal toxicity (Cobbett and Goldsbrough, 2002). They may also have evolved a range of mechanisms that enhance the uptake of metals from the soil solution and translocation within the plants (Reichman, 2001).
2.4.1 Tolerance mechanisms in germinating seeds Most studies on plant response to excess metals have used adult plants and seedlings (Cataldo et al., 1978, Gabrielli et al., 1991) or seeds of major food plant species, such as crops and pastures (Munzuroglu and Geckil, 2002). There have been few studies on the metal tolerance mechanisms of seeds of species from serpentine soils during germination. Leon et al. (2005) studied the effects of three Ni salts on germinating seeds of Grevillea exul var. rubiginosa (Figure 2-11), an endemic serpentine species of New Caledonia and proposed metal exclusion from seeds as a possible tolerance mechanism in this species. More specifically, Leon et al. (2005) observed that calcium crystals found in the seed coat of Grevillea exul var. rubiginosa seeds could induce the crystallization of Ni by a heterogenic nucleation mechanism that forms a barrier to Ni entry (Figure 2-12). They concluded that in this species, the seed coat which is rich in calcium crystals could play a role in excluding Ni and protecting the inner part of the endosperm against metal toxicity (Barnabas and Arnott, 1990, Leon et al., 2005).
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Figure 2-11 Surface view of a seed of Grevillea exul var. rubiginosa; a) seed coat; b) position of the embryo; c) endosperm area (Source: Leon et al., 2005).
Figure 2-12 SEM images and point analysis in the seed coat of a germinating seed of G. exul var. rubiginosa treated with 500 mg Ni L-1; (A) Longitudinal section of a seed: a) endosperm; b) seed coat with a hypodermal cell layer incrusted by crystals; (B) Detail of the inset area in A indicating crystal locations: a) outer layer of the seed coat; b) crystal in the seed coat; c) sclereid cell in the seed coat; d) inner layer of the seed coat (Source: Leon et al., 2005).
2.4.2 Tolerance mechanisms in established plants Adult plants exhibit a variety of metal tolerance mechanisms including metal chelation, transport and compartmentalization (Verbruggen et al., 2009). Most of the research on metal transport has been directed towards mechanisms of uptake and accumulation under conditions of deficiency (Marschner, 2012), or the characteristics of hyperaccumulation (Krämer, 2010). While the 64 molecular aspects of this field provide many opportunities for fundamental research (DalCorso et al., 2013) the following discussion will be restricted to tissue and organ scales of organization.
2.4.2.1 Restriction of uptake or transport Metal exclusion from the plant: excess metals are prevented from entering the plant either by precipitation or by complexing metals in the root environment therefore reducing their toxic effects (Khan et al., 2000, Reichman, 2002). For example, pH of the rhizosphere of Al tolerant mutant mouse ear cress (Arabidopsis thaliana) increased in response to excessive concentration of Al (Degenhardt et al., 1998) and exudation of citric acid from the roots of maize (Zea mays L.) increased at elevated Al concentrations (Pellet et al., 1995). An increase in carbon content of the rhizosphere of Lupinus albus roots may increase the availability of metals in neutral soils, but in acid soils, the concentration of CaCl2-extractable Cu, Zn and Mn were lower than in the bulk soil due to increased retention of Fe(III) hydroxides/oxyhydroxides (Martinez-Alcala et al., 2010).
Cellular exclusion: this mechanism avoids the build-up of excessive metal concentrations in the cytosol by their sequestration in the apoplastic free space of the roots, thus preventing the appearance of toxicity symptoms (Hall, 2002). Tice et al. (1992) observed more symplastic Al in the roots of a sensitive cultivar than in a tolerant cultivar of common wheat (Triticum aestivum L.), suggesting the existence of an exclusion mechanism. Colzi et al. (2011) suggested that copper tolerant individuals of Silene paradoxa did not take up copper because lower pectin concentrations in the cell walls reduced the number of copper-binding sites and the opportunities for transfer to the symplast.
Complexation at the cell wall-plasma membrane interface: metals such as Fe, Cu, Zn and Pb can accumulate at high concentrations at the cell wall-plasma membrane interface, a proposed site of metal tolerance, and they are found associated with Si contained in the cell wall (Neumann et al., 1997). Tolerant plants have high cell wall CEC (mediated by COO-) to complex metals and prevent their entry to cells (Cseh, 2002). Lead accumulated in the cell walls of shoots of sago pondweed (Potamogeton pectinatus L.) as the plasma membrane of the species acted as a barrier to the entry of Pb into the vacuoles (Sharpe and Denny, 1976). In contrast, Johnson and Singhal (2013) showed that the chelate, EDTA, increased the transport of Cu from the root cortex to the vascular cylinder and to shoots of Lolium perenne, while the addition of citric acid reduced this transport and led to retention of Cu in the root cortex. This evidence suggests that there may be multiple mechanisms that promote or inhibit metal uptake into the vascular system of the root.
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Active efflux: metals such as Fe, Cu, Zn, and Al may be excreted by roots and leaves (Delhaize et al., 1993a, Xu et al., 2009). Another possible mechanism is metal loading and shedding of older leaves (Luna et al., 1994, Ouelhadj et al., 2007, Reichman, 2002), as occurred in oat (Avena sativa L. cv Suregrain) in the presence of Cu (Luna et al., 1994). Cadmium can be excreted through the salt glands on the surface of the leaves of salt cedar (Tamarix smyrnensis) (Manousaki et al., 2008).
Reduction of influx: in arsenate tolerant clones of common velvet grass (Holcus lanatus), reduction in arsenate influx was attained via suppression of the high affinity phosphate/arsenate uptake system causing reduced influx of both phosphate and arsenate (Meharg and Macnair, 1994).
Nonuniform distribution and translocation within plant: translocation of metals to the shoot is reduced. For example, translocation of Mn and Fe to the shoot was reduced in wheat (Triticum aestivum L.) (Wheeler and Power, 1995) and northern black wattle (Acacia auriculaeformis) (Wheeler and Power, 1995, Zhang et al., 1998). Some cells in the roots of maize (Zea mays) accumulated Cu and died while others continued to function normally (Ouzounidou et al., 1995, Reichman, 2002).
2.4.2.2 Compartmentation and complexation with the cell Compartmentation within vacuoles: metals may be stored in vacuoles, where they can be detoxified and do not interfere with cell metabolism (Rengel, 1997, Leitenmaier and Küpper, 2013). For example, arsenate is reduced to arsenite in fronds of Chinese brake fern (Pteris vittata) and arsenite is subsequently stored in vacuoles (Wang et al., 2002). Arsenic hyperaccumulators form sulfur ligands during transport, but arsenate is predominantly bound by oxygen ligands and stored in the vacuoles of leaf epidermal cells (Leitenmaier and Küpper, 2013). Al can accumulate to high concentrations in vacuoles and has been found to be associated with phytate and Si, which reduced its toxicity (Vazquez et al., 1999, Reichman, 2002).
Complexing by metallothioneins: metals are bound by low molecular mass, cysteine rich, metal- binding proteins (metallothioneins) which act as detoxification molecules (Robinson, 1990, Merrifield et al., 2004, Hassinen et al., 2011). Rauser (1984) proposed that a Cu-binding metallothionein found in redtop (Agrostis gigantea) accounted for its metal tolerance.
Complexing by phytochelatins and glutathione: metals are bound by thiol-rich metal-binding peptides phytochelatins (PCs) and glutathione (GSH) which act as detoxification molecules (Rengel, 1997, Cobbett, 2000, Tomaszewska, 2002, Yang et al., 2005, Gupta et al., 2013). For example, it has been shown that arsenic was stored as AsIIItris-glutathione (shoot) and AsIII-tris-
66 thiolate complex (root) in Indian mustard (Brassica juncea) probably by glutathione or PCs (Pickering et al., 2000).
Complexing by organic acids: metals are bound by organic acids (such as malic, malonic, citric, oxalic, formic and succinic acids), become unavailable to the plant, and therefore are less toxic (Prasad, 1999). For example, aluminium stress in buckwheat (Fagopyrum esculentum) caused secretion of oxalic acid into the rhizosphere (Zheng et al., 1998). In wheat (Triticum aestivum L.), Delhaize et al. (1993b) found that organic anions (malate and citrate) were released from roots and bound and detoxified harmful Al3+ cations in the apoplast. Nickel was bound by citric acid in the sap of Sebertia acuminata (Sagner et al., 1997).
Complexing by organic and inorganic ligands: metals are bound by phytate, histidine, proline, phosphorus, and sulphate (Burke et al., 1990, Reichman, 2002, Leitenmaier and Küpper, 2013). For example, Al was bound by phytate in vacuoles of Al tolerant maize (Z. mays) (Vazquez et al., 1999), and Zn was bound by phytate in the roots of tussock grass (Deschampsia caespitosa) (van Steveninck et al., 1987).
Alterations of cellular metabolism: metal-sensitive metabolic processes are avoided by the activation of alternative non-metal sensitive pathways, or the metal sensitivity of enzyme activity is counteracted by increased enzyme syntheses (Verkleij and Schat, 1990, Gallego et al., 2012). For example, increased synthesis of malic enzyme was observed with increasing concentrations of Zn and Cd in the leaves of dwarf beans (Phaseolus vulgaris L.) (van Assche et al., 1988).
2.4.2.3 Volatilization of metals Volatilization is the process whereby elemental or combined forms of metals accumulate in the apoplast of leaves, vaporise and diffuse into the atmosphere, thereby resulting in metal detoxification (Padmavathiamma and Li, 2007). For example, inorganic Se is volatilized by plants and microorganisms (Pilon-Smits, 2005) and Hg is released through plant stomata (Greger, 1999). Volatilization of As as toxic trimethyl arsine has been demonstrated in fungi and microorganisms (Edvantoro et al., 2004). In marine algae, volatilization of As as dimethylarsenite has been suggested as a tolerance mechanism (Fitz and Wenzel, 2002). Several arsenite S- adenosylmethionine (SAM) methyltransferases have been recorded in cyanobacteria and algae (Ye et al., 2012) but higher plants are not known to methylate As. Where this occurs, methylated arsenical compounds in plants probably originate from microorganisms in soils and the rhizosphere (Ye et al., 2012).
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2.4.2.4 Role of mycorrhizae Mycorrhizae are mutualistic associations between certain soil fungi and the roots of the majority of plant species (Reddell and Milnes, 1992, Reichman, 2002, Zhang et al., 2018). Generally, the majority of plants growing under natural conditions have mycorrhizae (Khan et al., 2000). However, plants belonging to the families of Brassicaceae and Caryophyllaceae are generally known as non-mycorrhizal (Hildebrandt et al., 2007) although a Brassicaceae species from metalliferous sites, such as Thlaspi praecox Wulfen was found to be mycorrhizal (Vogel-Mikus et al., 2005, Hildebrandt et al., 2007). The same energetic resources may be required for the development of either mycorrhizal associations or metal hyperaccumulation, so these traits may appear to be in conflict, so the coincidence of these traits in a single species might be the exception rather than the rule (Audet, 2013).
Mycorrhizae play important roles in plant metal uptake (Smith and Read, 2010, Meier et al., 2012). They can facilitate plant uptake and transport of less mobile soil nutrients such as P (Smith et al., 2011), reduce plant uptake or translocation of Cu, Pb and Zn (Chen et al., 2005), or increase plant metal tolerance by reducing metal translocation to shoots (Wilkinson and Dickinson, 1995, Reichman, 2002, Zhang et al., 2018). For example, combined use of arbuscular mycorrhyzae (Glomus etunicatum) and nitrogen increased the shoot content and specific absorption rate (SAR) of P and Zn, and reduced the Al shoot content and Mn SAR in wheat (Triticum aestivum L.) grown on acidic soil. This is because arbuscular mycorrhizae can concomitantly bind, immobilize or increase the bioavailability of different elements (Cornejo et al., 2008, Gadd, 2010). Accumulation of Pb was 10% less in mycorrhizal than in non-mycorrhizal maize (Zea mays) plants (Chen et al., 2005). Mycorrhizae can also protect plant roots from metals by acting as an effective exclusion barrier or via metabolic processes, such as intracellular accumulation and extracellular precipitation of metals by metabolites (e.g. sulphites of phosphate) in fungal exudates (e.g. saprophytic fungi) (Khan et al., 2000). However, mycorrhizae may increase the uptake of As in Chinese brake fern (Pteris vittata L.) because its chemical properties are similar to those of P (Agely et al., 2005, Danh et al., 2014).
2.5 Phytoremediation Phytoremediation involves the use of plants to remove, transfer, inactivate, stabilize/immobilize and/or degrade contaminants in soil, sediment and water (Horne, 2000, Padmavathiamma and Li, 2007, Dhankher et al., 2011). Advantages of phytoremediation over techniques such as excavation to landfill, are that it is performed in situ, which decreases the amount of soil disturbance, reduces contaminant spread via air and water and does not require expensive equipment or highly specialized personnel (Ghosh and Singh, 2005). There are two remediation strategies for metal contaminated soils using plants: (1) removal (phytoextraction) and (2) containment 68
(phytostabilization) of metals (Mendez and Maier, 2008a). The selection of a particular method depends on the metal contaminants, the site conditions, the extent of clean-up required and the types of plants available (Marques et al., 2009).
2.5.1 Phytoextraction Phytoextraction is the use of metallophytes (i.e. metal accumulators and hyperaccumulators) to extract metals from the soil and concentrate them in shoots (Blaylock and Huang, 2000, Ghosh and Singh, 2005). Shoots are then harvested to remove contaminants from the sites for extraction, storage or safe disposal (Blaylock and Huang, 2000) by incineration, secured landfill (Kim and Owens, 2011) or ocean dumping (Zhang et al., 2011). Characteristics of plants suitable for phytoextraction are tolerance to a specific metal; ability to accumulate several metals in large amounts; adaptation to soil and climate; and root systems that conform to the spatial distribution of the pollutant (Padmavathiamma and Li, 2007). However, field crops (e.g. Zea mays, Phaseolus vulgaris and Sorghum bicolor) which are not hyperaccumulators have been investigated because of their high biomass production and dense patterns of shoot and root development (Vamerali et al., 2010). Phytoextraction involves repeated growing and harvesting of plants until the soil metal concentration and environmental risk of metal leaching are reduced to acceptable values (Kumar et al., 1995, Blaylock and Huang, 2000). This is a continuous and long-term process (1 to 20 years) (Salt et al., 1998, Vamerali et al., 2014). Phytoextraction is suitable for remediating large areas of land contaminated at shallow depth with low to moderate concentrations of metal contamination (Padmavathiamma and Li, 2007).
2.5.2 Phytostabilization Phytostabilization involves the use of metallophytes (i.e. metal excluders) to stabilize (rather than to remove) metals in soils by their retention in the soil or roots (McCutcheon and Schnoor, 2003). Metals are stabilized in the soil by absorption and accumulation in the roots, or are immobilized by adsorption onto roots or precipitation within the root zone and by physical stabilization of soils and reduction of metal leaching through erosion (Ghosh and Singh, 2005, Pilon-Smits, 2005).
Phytostabilization is mostly used for the remediation of soil, sediment and sludges (Ghosh and Singh, 2005) but it is also applicable to a wide range of sites, including large abandoned sites or large areas of surface contamination (Willscher et al., 2013), urban areas and fine-textured soils with high organic matter content (Berti and Cunningham, 2000). The method is particularly appropriate for elemental contaminants in soils where the best alternative is to physically hold or immobilize the pollutants in place to (1) prevent greater bioavailability during removal, (2) preserve ground and surface water, and (3) reduce potential exposure of human and wildlife to substances of 69 concern (McCutcheon and Schnoor, 2003). This technique does not require the disposal of biomass. However, it is not suitable for highly contaminated sites where it is impossible for plants to survive and grow (Padmavathiamma and Li, 2007). In addition, the contaminant remains and therefore requires regular monitoring (McCutcheon and Schnoor, 2003, Ghosh and Singh, 2005).
Plants appropriate for phytostabilization have particular characteristics including: tolerance to extreme concentrations of metals; accumulation of metals in the roots with no translocation from roots to shoot; and high root biomass production capable of immobilizing the contaminants in soil via accumulation in roots or precipitation within the root zone (Padmavathiamma and Li, 2007).
Native plant species are preferable for phytostabilization as they have adapted to local environmental stresses, such as drought, low pH and salt. Unlike exotic species, native species present the advantage of not being potentially invasive and they therefore preserve and possibly increase local biodiversity (Mendez and Maier, 2008a). Native grasses are often used for phytostabilization as they normally have dense canopies and root systems (Yang et al., 2014), are often tolerant to high metal concentrations (Baker et al., 1994b), and exhibit fast growth, massive root biomass, strong resistance to diseases and effective stabilization of soils (Koptsik, 2014). Therefore, grasses are excellent for rehabilitation of degraded mine lands as they physically stabilize the soil against wind or rain impact and erosion or leaching, and they restrict off-site migration of contaminants (Cunningham and Berti, 2000). However, most of the grasses used thus far for phytostabilisation are of European origin (Baker, 2007).
Australian native grasses are considered likely to establish successfully in mine sites and there has been an increasing demand for their use in land rehabilitation generally (Grice et al., 1995). Huxtable (2000) listed a range of Australian native grass species which have been used or could be potentially suitable for rehabilitation of contaminated mine sites including Mitchell grass (Astrebla spp.), spear grass (Austrostipa spp.) and Kangaroo grass (Themeda australis) among many others. More recently, Lottermoser et al. (2009) studied metal tolerance and uptake of Mitchell grasses on contaminated mine wastes. They concluded that curley Mitchell (Astrebla lappacea), bull Mitchell (Astrebla squarrosa), hoop Mitchell (Astrebla elymoides) and barley Mitchell (Astrebla pectinata) are tolerant of metals and have the ability to accumulate Pb and Zn into their shoots above the normal range for pasture species but below the foliar threshold for accumulators (see Table 2-4). Doronila et al. (2014) showed in glasshouse trials that Bothriochloa macra and Enteropogon acicularis, two warm season drought-tolerant Australian native grasses, could grow satisfactorily on As-rich waste rock, but much less so on saline tailings. Both species were classified as metal excluders. However, there is still very little information on the metal tolerance of Australian native 70 grasses and further investigation is needed of their suitability for phytoremediation purposes and their metal tolerance mechanisms.
2.6 Nanotechnology 2.6.1 Nano/microparticles and their applications in agriculture Nanotechnology is the engineering of functional systems at the nanoscale (1-100 nm) (Sun et al., 2006, Klaine et al., 2008, Hussain et al., 2009). It is an emerging and rapidly developing industry that has promising applications to the economy, society, and the environment nationally and internationally (Lin and Xing, 2007). Worldwide investment in nanotechnology by government organizations has increased approximately sevenfold from US$432 million in 1997 to almost US$3 billion in 2003 (DeFrancesco, 2003) and increased to more than US$6 billion by 2009 (Brumfiel, 2003).
Engineered nanomaterials can be classified into four broad categories (Lin and Xing, 2007): (1) Metal-based materials such as quantum dots, nanogold, nanozinc (nano-Zn), nanoaluminium
(nano-Al), and nanoscale metal oxides like TiO2, ZnO and Al2O3; (2) carbon-based materials, usually including fullerene, single walled carbon nanotubes (SWCNT) and multiwalled carbon nanotubes (MWCNT); (3) dendrimers, which are nano-sized polymers built from branched units capable of being tailored to perform specific chemical functions; and (4) composites, which combine nanoparticles with other nanoparticles or with larger, bulk-type materials.
The most obvious advantage that nanostructured materials provide for environmental remediation is that they offer very high specific surface area (measured in square meters per gram) (Fryxell and Mattidod, 2006). These true nano size effects are due to the proportion of surface sites at edges or corners, the presence of distorted high energy sites, contributions of interfacial free energies to chemical thermodynamics, the effects of altered surface regions, and quantum effects (Giammar et al., 2007). Nanomaterials have well defined structure, high reactivity, are easily dispersed, and are readily tailored for application in different environments, including chemicals or materials developed for the detection and remediation of pollutants in the environment (Khin et al., 2012, Trujillo-Reyes et al., 2014, Zhang et al., 2015).
Water soluble polymers (polyacrylamides) have been used as soil conditioners in agricultural areas and appear to have various beneficial soil amendment properties, including minimization of water run-off, erosion and crusting, and stabilization of soil structure by preventing clay dispersion
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(Barvenik, 1994, Shainberg and Levy, 1994). Significant erosion reduction results when small amounts of polyacrylamide are dissolved in irrigation water and deposited on the soil surface (Bicerano, 1994). Polyacrylamides can be expected to be slowly lost from soil over time as a result of mechanical degradation, chemical and biological hydrolysis, sunlight, salt and temperature effects (Wallace et al., 1986, Tolstikh et al., 1992) with an estimated half life of about 10 years (Azzam et al., 1983) which was confirmed by stable isotope analysis (Entry et al., 2008).
2.6.2 Nano/microparticles and metals Because of their high surface area, nanomaterials can be ideal for metal sequestration (Huang et al., 2017). For environmental applications, nanotechnology offers the potential of novel functional materials, processes and devices with unique activities towards unmanageable pollutants (Sun et al., 2006, Guiwei et al., 2008). Nanotechnology is seen as the generation of remediation technology which will improve or replace conventional techniques (Sun et al., 2006). This type of nanotechnology is very promising for cost effective in situ remediation. For example, Bell et al. (2006) studied the synthesis of surface functionalised polymer nanoparticles for the selective and irreversible sequestering of metals such as Hg, Pb, and As in water. Nolan et al. (2006), Kara et al. (2004), Kesensci et al. (2002) and Xu and Zhao (2007) have also shown the successful use of polymer nanoparticles for remediation of metal-contaminated water.
Other types of nanoparticles have been studied, including vivianite and dendritic polymers and iron phosphate (vivianite) nanoparticles (Liu and Zhao, 2007). Liu and Zhao (2007) observed that nano- sized vivianite reduced the leachability and bioaccessibility of Cu(II) in soil (calcareous, neutral, and acidic). Dendritic polymers can remove all the exchangeable Cu(II) from a water solution Cu(II) (Vliet et al., 2007). Li and Zhang (2007) showed that iron nanoparticles can reduce metals to insoluble forms, immobilize them in the soil and prevent food chain contamination. Xu and Zhao (2007) reported that engineered stabilized zero-valent iron (ZVI) nanoparticles were able to immobilize Cr(VI) in water and in a sandy loam soil.
In vitro and in vivo tests on the capacities of different crosslinked acrylamide hydrogel polymers (non-functional, xanthate functional and thiol functional) of different particle size (micro- and nano- sized) to bind Pb, Cu, As and Zn in solution were studied by Rossato et al. (2009). These authors found that thiol functional polymers reduced the total soluble concentrations of Pb (9650 µM), Cu (4000 µM) and Zn (1000 µM) by 86.5, 75.5 and 64%, respectively, whereas, they excluded As (667 µM). The particles tested in this study exhibited a high water holding capacity from 470 – 1,060% of their dry mass. Furthermore, the thiol functional polymer (X3) was not toxic to seed germination and allowed normal germination and root elongation of the native grass curly Mitchell grass 72
(Astrebla lappacea) at available metal concentrations (Pb 4825 µM and Zn 10000 µM) normally phytotoxic for that species (Rossato et al., 2011).
To date, the metal-binding and water-holding capacities of X3 have never been tested in metal- contaminated substrates.
2.6.3 Nano/microparticles and the environment Nanotechnology is being used increasingly in many areas, including environmental management. For example, polyacrylamide based polymers used in agriculture degrade in soil and aquatic environments and their decomposition products are not toxic to the environment (Barvenik, 1994, Seybold, 1994) or to humans, animals, fish and plants (Seybold, 1994). Polyacrylamide used as a soil conditioner is unlikely to pollute the soil with sufficient acrylamide monomer (which is believed to be toxic) arising from depolymerisation to constitute a potential hazard (Friedman and Rasooly, 2013). Analysis of residual acrylamide in beans, corn, potatoes, and sugar beets grown in soil treated with polyacrylamide to reduce erosion showed concentrations of <10 ppb (Friedman, 2003).
Despite all the evidence available, the application of nanotechnology is still of great concern, including the potential of nanomaterials in the environment for adsorption and assisted transfer of toxic substances in water, air and soil, or living organisms, biodegradability and persistence of polymer-based and other relevant nanomaterials, and the potential sources of releasing toxic nanomaterials into the environment (Dionysiou, 2004, Brar et al., 2010, Canesi et al., 2010). Lee et al. (2008) found that the addition of Cu nanoparticles reduced the growth rate of mung bean (Phaseolus radiatus) and wheat (Triticum aestivum) seedlings cultivated on agar. However, there is still very little information on whether nanoparticles may become toxic with time or on their stability and durability in relation to metal sequestration (Vliet et al., 2007). Particle size is an important characteristic for the transport of materials in porous media and their small size means that nanoparticles may not remain at the point of application in the environment (Hussain et al., 2009). Therefore, further research is needed to reveal the transport and fate of nanoparticles in the environment and the risks of toxicological impacts on living organisms, so that the technology can be used in a safe and conservative manner for the remediation of metal contaminated sites.
2.7 Hypothesis The following hypothesis will be tested in this thesis:
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The addition of micron-sized metal-binding particles to metal-contaminated soils and mine waste materials prior to revegetation can increase the certainty and speed of vegetation establishment on these substrates by: Reducing the plant available concentrations of metals in the substrates, and by Increasing the water holding capacity of the substrate and enhancing its supply to plants.
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Chapter 3. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters
The chapter has been accepted and published as a research article in a high ranking refereed journal, Journal of Hazardous Materials, 2019 with the following detail:
Guterres, J., Rossato, L., Doley, D., Pudmenzky, A. 2019. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters. Journal of Hazardous Materials, Vol. 364, p.449-467.
For uniformity of the thesis, some changes in the formatting have been made including omitting the ‘article highlights’ and inclusion of the references in the complete thesis reference list.
Results from some parts of the data collected for the chapter were published in the Proceedings of the Seventh International Conference on Serpentine Ecology, 12-16 June 2011, Coimbra, Portugal, under the title ‘A geobotanical survey of serpentine and metalliferous mine sites in Australia led to the discovery of new native metallophyte species and their metal-accumulation characteristics’.
Field data collections and plant species identification were performed by Dr Don Butler and Ms Joy Brushe from the Queensland Herbarium (Brisbane, Australia), as well as by staff from mine sites and Dr Laurence Rossato and Dr Alex Pudmenzky. Laboratory analyses of plant samples were conducted by the candidature under the supervision of Dr Laurence Rossato, Dr David Doley and Dr Alex Pudmenzky with the assistance of Mr David Appleton and Mr Steve Appleton from the School of Agriculture and Food Sciences’ analytical services at the University of Queensland.
Data analysis and preparation of figures described in the paper were performed by the candidate and under the supervision of the advisors Dr Laurence Rossato and Dr David Doley, and Dr Alex Pudmenzky.
3.1 Abstract Based on a review of the literature, we have developed a functional conceptual framework of plant metal uptake in relation to plant available metal concentration in the soil. This framework applies to all plant parts and plant available metal concentrations in soils, and was validated using independent datasets from field surveys and the literature.
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This is the first framework based on metal transporter kinetic parameters and combining Michaelis- Menten (hyperbolic) kinetics facilitated by the High Affinity Transport System (HATS) for soil concentrations below the transition concentration between transport systems, and linear metal uptake facilitated by the Low Affinity Transport System (LATS) for higher soil available metal concentrations.
We propose a new terminology for metal tolerant plants, i.e. metal tolerators, based on this framework. Depending on the plant available metal concentrations in the soil, tolerator responses to metals can be described best by either Vmax and Km for soil concentrations below the transition concentration between metal transport systems (HATS), or by the slope for greater soil concentrations (LATS).
This conceptual framework may be a useful tool for selecting suitable metal tolerators for specific phytoremediation purposes, and may be also applied to non-metal elements or ions.
Keywords: metal transporter kinetics, HATS, LATS, plant metal responses, metal tolerator
3.2 Introduction In order to maintain healthy metabolism, growth and development, all organisms, including plants, need large amounts of macronutrient elements such as nitrogen, phosphate, potassium, sulfur, calcium and magnesium, and small quantities of many essential micronutrient ions including zinc, nickel, copper, manganese, iron, molybdenum, boron and chloride (Marschner, 2011). Excessive amounts, especially of these essential metal micronutrients, can cause symptoms such as abnormal growth, chlorosis and necrotic spotting and decrease plant growth or quality (Bennett, 1993, Kabata-Pendias, 2001, Marschner, 2011). On the other hand, non-essential metals such as cadmium, lead and mercury and the metalloid, arsenic, are toxic at lower concentrations due to their disruption of enzyme functions, replacement of essential metals in pigments or generation of reactive oxygen species (Babula et al., 2009).
The survival of plants in environments with either very low or elevated concentrations of potentially toxic elements has been a topic of interest to fundamental biologists (e.g., Baker and Proctor (1990); Bennett (1993)) and also to environmental managers (Mendez and Maier, 2008a). Therefore, it is important to have a clear and consistent means of describing the responses of different plant species to different environmental conditions.
Similar to other organisms, plants have evolved various mechanisms that maintain physiological concentrations of essential metal ions and reduce the effects of non-essential metals (Meharg, 2005,
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Manara, 2012). Metal homeostasis is attained via regulation of ion uptake and efflux, detoxification via complexation or transformation, intracellular compartmentation, such as vacuolar storage, and intercellular transport to specialized cells (Meharg, 2005, Krämer et al., 2007, Das et al., 2011, Andresen et al., 2018), or by excretion (Weis and Weis, 2004, Maathuis et al., 2014). Homeostasis depends on metal ion transport from the apoplast to the symplasm, mediated by membrane proteins with transport functions, known as transporters, as the ions cannot move freely across the lipophilic cellular membranes (Clarkson, 1993b, Lasat, 2000, Manara, 2012, Maathuis et al., 2014).
Uptake of metal ions by plants may be facilitated by four different types of transport systems depending on the external concentration of metal ions (Fraisier et al., 2000, Laugier et al., 2012, Nikolic et al., 2012). At low external ion concentrations, saturable ion transport systems, constitutive high affinity transport systems (CHATS) and inducible high affinity transport systems (IHATS) may be involved in ion influx (Dechorgnat et al., 2011, Sorgonà et al., 2011, Nikolic et al., 2012). For example, for nitrate, CHATS is constitutively expressed, has a high affinity for nitrate and transports at a low rate (Faure-Rabasse et al., 2002), whereas IHATS is induced by the presence of nitrate and has a higher uptake capacity (Sorgonà et al., 2011). Puig (2014) concluded that, among a family of high-affinity Cu transport (COPT) proteins, COPT1 facilitates Cu uptake by roots, COPT6 enables Cu distribution to the shoot, and COPT5 activates Cu mobilization from storage organelles. Abdin et al. (2011) observed that sulfate uptake of Brassica juncea cv. Pusa Jai Kisan was improved by constitutive over-expression of a Lycopersicon esculentum sulfate transporter (LeST 1.1.) gene which encodes a high-affinity sulfur transporter in the root epidermis. Vert (2002) and Connolly et al. (2002) identified an iron transporter (IRT1) as an IHATS system responsible for high-affinity Fe uptake in Arabidopsis thaliana under iron deficiency conditions. Another IHATS transporter is NRAMP1 (Natural Resistance Associated Macrophage Protein 1), the main high-affinity Mn transporter in Arabidopsis which is stimulated by Mn deficiency (Cailliatte et al., 2010).
At high external ion concentrations, non-saturable constitutive low affinity transport systems (CLATS) and inducible low affinity transport systems (ILATS) may facilitate ion transport (Nikolic et al., 2012). Wei et al. (2009) identified the ILATS iron transporter TcNRAMP3 in Thlaspi caerulescens where it was induced by Fe-starvation and by the presence of the heavy metals Cd and - Ni in the growth medium. Nicolic et al. (2012) observed in Arabidopsis that the NO3 transporter NRT1.1. has an essential role in ILATS functions.
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Under nutrient sufficient conditions, constitutive absorption systems (CHATS and CLATS) play a key role in nutrient uptake, but under nutrient deficiency, inducible systems (IHATS and ILATS) increase the rate of transport of a particular nutrient (Graham and Stangoulis, 2003).
Numerous studies have shown that, in both plants and animals, metal uptake and transport by high affinity metal transport systems (HATS) generally follow Michaelis-Menten kinetics. Michaelis- Menten kinetics describe the rate of enzymatic reactions (V) as a function of the concentration of a substrate, [S], the maximum reaction rate reached by the system at saturating substrate concentration, Vmax, and the Michaelis constant, Km, which is the substrate concentration at which the reaction rate is half of Vmax (Down and Riggs, 1965, Ritchie and Prvan, 1996).
V = Vmax.[S]/(Km +[S]).
The smaller the Km value, the higher the affinity of a transporter for the substrate, which means it needs less substrate to achieve half of Vmax and the enzyme is a more effective catalyst for the reaction (Shiflet and Shiflet, 2006). The transport of macronutrients is well described by this relationship. For example, Cerezo et al. (2010) identified a high affinity nitrate transport system - (HATS) which displays Michaelis–Menten kinetics functions at [NO3 ] lower than 1000 µM. The - HATS appeared to be substrate inducible (IHATS) since NO3 influx as well as net uptake rate - - increased substantially when NO3 was supplied after a period of NO3 starvation (Cerezo et al., 2010).
Michaelis-Menten kinetics also apply to micronutrient elements. Pedas et al. (2005) showed that kinetic parameters for high-affinity root Mn2+ influx in 10-day-old plants of the Mn-efficient (the genotype which effectively uses Mn to produce high yield) barley genotype Vanessa (Vmax= 0.0054±0.0010 µmol Mn g root DW-1 h-1) exceeded those of the Mn-inefficient genotype Antonia -1 -1 (Vmax= 0.0014±0.0004 µmol Mn g root DW h ) at external Mn concentrations below 0.13 µM, although there was no significant difference in Km (0.0054±0.0018 and 0.0027±0.0016 µM, respectively).
The metalloid As is a P-analog, and enters plant roots as arsenate As(V) on a phosphate-transport protein (Meharg and Macnair, 1992a, Poynton et al., 2004, Danh et al., 2014). As(V) influx can be described by Michaelis-Menten kinetics and the parameters Vmax and Km were lower in the As hyperaccumulating species Chinese brake fern (Pteris vittata L.) (0.1200±0.0180 µmol As g root DW-1 h-1 and 1.1000±0.3000 µM, respectively) than in a non-accumulating fern (Nephrolepis exaltata L.) (0.144±0.0200 µmol As g root DW-1 h-1 and 9.9000±2.0000 µM, respectively),
78 indicating higher affinity of the transport protein for arsenate, although phosphate inhibited arsenate influx in a directly competitive manner (Poynton et al., 2004).
Cohen et al. (1998) also demonstrated Cd2+ influx in roots of Fe-deficient pea (Pisum sativum L. cv. -1 -1 Sparkle) seedlings displaying Vmax around 2.36±0.25 μmol Cd g root DW h and Km around
1.5±0.6 μM, and in roots of Fe-sufficient pea seedlings exhibiting Vmax around 0.34±0.08 μmol Cd -1 -1 g root DW h and Km around 0.6±0.09 μM.
Similarly, Lasat et al. (1996) also observed a saturable component of Zn influx in the Zn -1 -1 hyperaccumulator Thlaspi caerulescens (Vmax of 2.7 μmol Zn g root DW h and Km of 8 μM) and -1 -1 the non-accumulator T. arvense (Vmax of 0.6 μmol Zn g root DW h and Km of 6 μM) that followed Michaelis-Menten kinetics.
At higher external concentrations, the rate of ion uptake in plants normally increases linearly with increasing substrate concentrations (Faure-Rabasse et al., 2002, Meng et al., 2014). Cerezo et al. (2010) stated that a constitutive low affinity transport system (CLATS), which displayed linear - kinetics, plays a major role in the uptake rate of the macronutrient NO3 at external concentrations higher than 1000 µM. Similar conditions have also been observed by Cerezo et al. (2007) in - Cleopatra mandarin and Troyer citrange seedlings at NO3 concentrations between 1000 and 10,000 µM.
Pedas et al. (2005) observed that, over the low affinity concentration range (>0.13 μM), influx of the essential micronutrient Mn was similar between Mn-efficient (Vanessa) (regression intercept at 2.0440 µmole Mn g root DW-1 h-1 and slope of 0.0105 µmole Mn g root DW-1 h-1 µM -1) and Mn- inefficient (Antonia) (regression intercept at 1.6607 µmole Mn g root DW-1 h-1 and slope of 0.0081 µmole Mn g root DW-1 h-1 µM -1) 10-day old barley genotypes, showing linear kinetics (calculated from authors’ data). Pedas et al. (2005) concluded that the uptake rates of Mn by LATS would certainly cause Mn toxicity if sufficient Mn was present for a few hours and if there was no system to facilitate efflux. Cohen et al. (1998) also demonstrated linear Cd influx kinetics of Fe-deficient and Fe-sufficient pea seedlings over a low affinity concentration range (>10 µM), and both sets of seedlings showed similar slope values of 0.0670±0.0035 µmole Cd g root DW-1 h-1 µM -1 for Fe- deficient roots and 0.0530±0.0013 µmole Cd g root DW-1 h-1 µM -1 for Fe-sufficient roots, but different regression intercept values of 2.55 and 0.30 µmole Cd g root DW-1 h-1 (calculated from authors’ data), respectively.
At concentrations >100 μM, Meharg and Macnair (1992a) showed linear trends of As influx in roots of tolerant and non-tolerant genotypes of Holcus lanatus L. with similar slope values of 79
0.0003 and 0.0002 µmole As g root DW-1 h-1 µM-1, respectively (calculated from authors’ data). However, the intercept values were significantly different, 0.1660 and 1.4017 μmole As g root DW- 1 h-1, respectively (calculated from authors’ data) (Meharg and Macnair, 1992a).
Metal ion uptake kinetics are typically determined after exposure of plants to experimental conditions for several days (Santa-Marıa and Cogliatti, 1998, Pedas et al., 2005, Cailliatte et al., 2010). It is hypothesized that plants grown in the field on substrates with different available metal ion concentrations will show similar patterns of ion uptake over an entire growing season and similar ion transporter kinetics as do plants grown in experimental systems for shorter periods of time. It is proposed that available metal concentration in the soil varies with soil water content but that it can be integrated over time in the same way that metal concentration in a plant part can represent the integral of uptake rates over a growing period. It is proposed that, at the level of the whole plant part and the bulk soil sample, each component can be regarded as an internally uniform system for the purpose of explaining material transfers at its boundaries.
Based on this review of the literature, the aims of the present paper were: i) to identify new key parameters for the characterization of metal-tolerant plants (metallophytes) that might have been omitted from current metallophyte classifications (Wei et al., 2009, Cerezo et al., 2010), and ii) to propose a new improved metallophyte classification incorporating these metal uptake parameters. In the absence of direct information on metal transporter parameters for the vast majority of native plant species anywhere in the world, it is important to establish from field collections of plant material whether their quasi-equilibrium concentrations of metals in different plant parts can be used to indicate the order of magnitudes of these metal uptake parameters. This latter objective will be tested initially on native plant species in Queensland, Australia.
3.3 Materials and methods 3.3.1 Plant and soil sample collection and treatment Plant and soil samples were collected from serpentine and metal contaminated mine sites in Queensland (Australia). Thirty-eight plant species were collected from Ridgelands & Rockhampton and Marlborough districts in Central Queensland (the largest serpentine area in Australia), 2 plant species from Mt. Morgan (an abandoned gold mine) and 3 plant species from Cracow (an operating gold mine) mine sites (Table 3-1 and Appendix I) between 2007 and 2008. Overall, 43 plant species were collected including grasses, trees and shrubs belonging to the Cyperaceae (4), Mimosaceae (7), Myrtaceae (12), Poaceae (17) and Thymelaeaceae (3) families.
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Table 3-1 Summary of plant and soil sample identification used in this study. Areas/Sites Site Sample ID Ridgelands, 1 1/2, 1/3a, 1/3b, 1/4, 1/5a, 1/5b, 1/6, 1/7, 1/10, 1/11, 1/12 Rockhampton & 2 2/1, 2/2, 2/4, 2/7, 2/10 Marlborough 3 3/1 4 4/3, 4/5, 4/7, 4/8, 4/9 5 5/6, 5/9, 5/10, 5/11, 5/12, 5/15, 5/18 6 6/3 7 7/2 8 8/1, 8/4, 8/5, 8/6, 8/7, 8/8, 8/9, 8/11, 8/12 Mt. Morgan mine site Mor Mor-1, Mor-5 Cracow mine site WTP WTP237, WTP239, WTP243
For each plant sample, fresh mature leaves (referred to as leaves henceforward) and, when accessible, roots which belonged to the same plant and senescent leaves were placed between layers of clean and metal-free-paper for drying and transport to the laboratory. The samples were then air dried, wiped clean of dust and separated into young or older leaves, senescent leaves and roots. All samples were cut into small pieces.
Representative composite surface soil samples (top 10 cm) were collected in close proximity to the plant species. Each composite soil sample was placed into a carefully labelled plastic zip-lock bag for transport to the laboratory for further treatment. The soil samples were then air dried at room temperature until constant weights were reached.
Sampling GPS coordinates were recorded for each plant species collected, and plant species were identified by staff from the Queensland Herbarium (Brisbane, Australia).
3.3.2 Soil analysis Each air-dried composite soil sample was ground with a mortar and pestle and passed through a 2- mm polyethylene sieve. The samples were then analysed for total soil metals and soil plant available metals.
For total metal determinations, each composite soil sample was reduced to fine particles using a ball mill (Planetary Ball Mills PM 200, RETSCH, Germany) at 650 rpm for 1 min. Then, 0.15 - 0.2 g of each soil sample was weighed into three acid digestion tubes to form three replicates. Metal-free acids (5 mL nitric acid 70% from Labscan Asia Co. Ltd., Bangkok, 2 ml hydrochloric acid 32% and 2.5 mL hydrofluoric acid 50% from Ajax Finechem, Australia) were added to each sample replicate 81 which was then heated in a microwave digester (MDS-200, CEM Corporation) at 120 atm (12.16 MPa) and 185 ˚C for 30 minutes. Each replicate of digested sample was transferred into 50-mL polypropylene conical tubes and boric acid ≥99.5% (ACS Reagent, Sigma Aldrich) was added as needed to minimize flocculation of the solution. The samples were then diluted with triple deionised water (TDI) up to either 30 or 35 mL. The samples were shaken well and then 10 mL of diluted samples were transferred into 10-mL Polypropylene Transparent (PPTR) tubes for inductively coupled plasma optical emission spectroscopy (ICP-OES) analysis of Al, Ca, Cu, Fe, K, Mg, Mn, Na, P, S, and Zn. The samples were diluted a further 10 times prior to inductively coupled plasma mass spectrometry (ICP-MS) analysis of Hg, Cr, Co, Ni, As, Se, Cd, Pb, Mo and Au. The standard soil reference material (SRM) (2709 San Joaquin soil) from National Institute of Standards and Technology (NIST), Gaithersburg (USA), provided by the School of Agriculture and Food Sciences (SAFS), the University of Queensland (UQ), Australia, was used for quality assurance and quality control protocol. The SRM 2709 San Joaquin soil contained certified values for most metals, particularly Al, Cu, Fe, Mn, Ni and Zn and was homogeneous. For ICP analysis, 0.15 - 0.2 g of the NIST SRM 2709 was weighed into three acid digestion tubes to create three replicates. Handling of the NIST was done according to quality control/quality assurance laboratory procedures for SAFS. The NIST and blanks were incorporated into the scheme of analysis from the beginning of sample digestion and throughout the entire analytical procedure under the same conditions as the samples. Separate aliquots of blanks, the NIST and samples were prepared for each ICP measurement (ICPMS or ICPOES). Three blanks and three NIST samples were included for every batch of up to 60 samples. All the analytical results of the NIST and samples were blank corrected. The NIST recovery fell within 80-120% of certified values, the acceptable standard (Klesta and Bartz, 1996), thus, the analytical results were considered to be of an acceptable quality. In some cases where percent recovery fell outside the standard range, the analytical results were considered to be biased. In such cases, all the samples within the same batch of this particular NIST were rerun.
Plant availability of metals in soils was measured via 0.01 M calcium chloride (CaCl2) extraction using CaCl2.2H2O (Ajax Finechem, Australia) (Menzies et al., 2007). Four grams of each composite soil sample (<2mm diameter size) were added to three polypropylene conical tubes to produce three replicates, followed by 40 mL of 0.01M CaCl2 solution added to produce a 1:10 soil : 0.01M CaCl2 suspension. The samples were placed in a Heidolph ReAx shaker at 40 rpm and room temperature for 3 hours to extract metals, then centrifuged for 5 minutes at 4000 rpm. After centrifugation, 10 mL of the solutions were transferred in a 10-mL PPTR tube and 250 µL of 70% concentrated nitric acid were added to each sample in order to maintain low pH and metal solubility. The samples were stored below 4˚C until analysis of plant available metals via ICP-OES was performed.
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3.3.3 Plant analysis For all plant samples, three replicates of 0.15 – 0.2 g of each plant part were weighed into three acid digestion tubes. Samples were then digested and analyzed for total metal concentrations following the procedure described for soil samples (in section 3.3.2). Samples were then heated in a microwave digester, transferred into a 50 mL polypropylene conical tube and boric acid was added to the solution in a similar way as for soil samples. Samples were then diluted with triple deionised water (TDI) up to 35 mL and shaken well. Thirty milliliters of the solution was transferred into 10- mL PPTR tubes for ICP-OES analysis of metals. A sample of NIST Certified Plant Material (1547 Peach leaves) was used for quality assurance and quality control. The NIST sample (from Gaithersburg, USA), provided by SAFS (UQ, Australia), had certified values for most of the metals analysed in this study, especially Al, Cu, Fe, Mn, Ni and Zn. Treatment for plant NIST followed the same procedure as for the plant samples. The quality control assurance protocol for plant NIST was identical to that of the soil NIST (see section 3.3.2 above).
3.3.4 Assessing possible plant sample contamination Plant samples taken for chemical analysis might be contaminated in two ways: i) physical or surface contamination of samples which arises before, during and after sampling but before the analytical process begins; and ii) chemical or laboratory contamination which can occur during sample analysis (Mitchell, 1960).
3.3.4.1 Physical or surface contamination of plant samples There is always the possibility of surface contamination of plants samples by fine dust from soil or other sources which contain very much higher concentrations of many elements than do the plants themselves (Mitchell, 1960, Markert, 1992, Wyttenbach and Tobler, 2002, Ferrari et al., 2006). In our study, all plant samples were subjected to surface decontamination prior to analysis. Leaf, stem and root surfaces were gently but thoroughly wiped with clean dry cellulose tissues (Kimwipes, Kimberly-Clark) to remove any residual materials including dust and soil particles. Samples were not washed as this is both ineffective for soil removal (Jones, 2001, Cook et al., 2009) and it can either leach target trace elements from plant tissue (Jones, 2001, Cook et al., 2009) or add elements to the tissue (Jones, 2001), further distorting the estimated tissue concentrations (Cook et al., 2009). For example, washing Citrus leaves with aqueous solution of 0.1% v/v detergent followed by a rinse in deionized water caused significant reductions of 67.2%, 48.4% and 44.7% of Fe, Mn and Zn, respectively, and moderate reductions of 8.9% of Cu (Labanauskas, 1968, Jones, 2001).
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3.3.4.2 Chemical or laboratory contamination During chemical and laboratory analysis, measurable quantities of elements or compounds being analysed may be introduced to the sample from various external sources, including from the laboratory atmosphere, the reagents and the apparatus used, and the analyst performing the analysis (Mizuike and Pinta, 1978, Ye et al., 2013). In our study, samples were prepared in a clean air- conditioned laboratory environment using non-metallic equipment to avoid metal contamination, such as clean glassware and polyethylene tubes. Reagents used were metal-free analytical grade nitric acid 70% (Labscan Asia Co. Ltd., Bangkok), hydrochloric acid 32% (Ajax Finechem Pty Ltd., Australia) and hydrochloric acid 50% (Ajax Finechem Pty Ltd., Australia). Blank correction was used to accurately determine the total amount of each element in a sample (Mitchell, 1960). Standard Reference Materials (SRM) from the National Institute of Standards and Technology (NIST) were also used to evaluate the reliability of analytical methods (Jones, 2001). By using the blank and the reference standards to verify analytical results, it is considered that any contamination that might have been introduced during laboratory analysis was eliminated (Jones, 2001).
3.3.4.3 Indicators of residual soil on plant samples Various indicator elements have been used in the literature to monitor possible soil contamination of plant samples on the basis that these elements are abundant in soil but not in plants (Cook, 2007, Cook et al., 2009). Therefore, sensitive indicators should present high soil/plant ratios, although precautions must be taken when using these ratios on samples of metallophytes taken from metal rich areas such as metalliferous mine sites, ultrabasic, ultramafic and serpentine geological areas (Mitchell, 1960). Indicators that have been used previously to assess possible soil contamination include titanium (Ti) (Mayland and Sneva, 1982, Cherney and Robinson, 1983, Cary et al., 1986, Cook et al., 2009, Fernando et al., 2009, Stangoulis, 2010), chromium (Cr) (Mitchell, 1960), aluminium (Al) (Cherney and Robinson, 1983, Cary et al., 1986), iron (Fe) (Mayland and Sneva, 1982, Cary et al., 1986, Fernando et al., 2009), silicon (Si) (Metson et al., 1979) and scandium (Sc) (Ferrari et al., 2006).
Titanium has been used most extensively as a soil contamination indicator for plant samples (Cherney and Robinson, 1983, Cary et al., 1986, Cook et al., 2009, Stangoulis, 2010) because of its low concentration (less than 3 mg.kg-1) in plants and the large (10,000 to 500,000) soil : plant concentration ratio (Mitchell, 1960, Mayland and Sneva, 1982). In addition, Cook et al. (2009) stated that Ti was considered a valid soil contamination indicator because there are three sources of Ti associated with plant samples, namely, contamination during laboratory processing, vascular uptake via roots and field soil residue on shoot surfaces. These authors surveyed soil and plant caesium (Cs) using Ti as soil contamination indicator and found that the concentration of Ti in leaf 84 samples was likely to be increased by 4 mg.kg-1DW through milling and by another 5 to 6 mg.kg- 1DW from reagents and labware. Root uptake accounted for 5 mg.kg-1DW of Ti in Crepis acuminata Nutt. seedling shoot tissues while dust applied to the shoot surface resulted in an eleven- fold increase in Ti concentration in the leaf samples. Cook (2007) and Cook et al. (2009) observed a strong relationship between Cs and Ti in shoot samples, and that higher Cs concentration in plants is caused by soil contamination and not uptake, suggesting Ti is a good indicator of soil contamination. Mayland and Sneva (1982) used a Ti dilution technique to determine the effect of soil contamination on the mineral composition of forage fertilized with nitrogen and found that iron (Fe) concentrations were unexpectedly high and were presumed to be associated with dust on the leaves. Although Ti is present in soils, its accumulation in leaf tissues is uncommon, and leaf concentrations exceeding 100 mg.kg-1DW are likely to be due to soil contamination (Fernando et al., 2009). Titanium has also been used as a soil contamination indicator in micronutrient analyses of wheat and rice (Stangoulis, 2010), and pasture (Metson et al., 1979). However, the use of Ti as a soil contamination indicator is valid only if the analytical sensitivity is similar for each of the elements (Mayland and Sneva, 1982). The concentration of Ti may vary with soil particle size and extraction or solubilisation of soil and plant samples should use the same procedures to account for similar solubilities in the soil and dust fractions (Mayland and Sneva, 1982).
Chromium (Cr) is one of the most abundant metals in the Earth’s crust (Panda and Choudhury, 2005, Paiva et al., 2009, Sundaramoorthy et al., 2010) with an average concentration of 100 mg Cr.kg-1 in soil (Kabata-Pendias and Mukherjee, 2007). In ultramafic rocks, Cr concentration can often exceed 1,000 mg.kg-1 and sometimes 100,000 mg.kg-1 (Kabata-Pendias and Mukherjee, 2007). In plants, Cr occurs mainly in roots (Cary and Kubota, 1990, Kabata-Pendias and Mukherjee, 2007), with shoot/root concentration ratios about 0.01 (Cary and Kubota, 1990, Sharma et al., 2005) and for both Cr(III) and Cr(IV) species in vegetables (Zayed et al., 1998). Leaves of plants grown on low-Cr soils generally had Cr concentrations less than 0.2 mg.kg-1 DW and often less than 0.1 mg.kg-1 DW (Cary and Kubota, 1990). Unsurprisingly, plants grown on high-Cr soils had higher Cr concentrations, although some of this could be attributed to soil contamination (Cary and Kubota, 1990). Under normal conditions, Cr concentration in plants is less than 1 mg.kg-1 DW (Oliveira, 2012) and it is considered to become toxic between 5 and 30 mg.kg-1 DW (Kabata-Pendias and Mukherjee, 2007).
Aluminium (Al) indicates possible soil contamination in plant samples (Metson et al., 1979, Stangoulis, 2010). Ranging between 1 and 4% in soil (Kabata-Pendias and Mukherjee, 2007), it is more abundant than Fe (Yasmin et al., 2014). However, because Al is not readily taken up by plants (Yasmin et al., 2014), it can be used as an indicator of soil contamination for most plants (Mitchell, 85
1960, Mayland and Sneva, 1982, Stangoulis, 2010, Yasmin et al., 2014). A high concentration of Fe with a correspondingly high concentration of Al (or Ti) in food crops may indicate contamination (Stangoulis and Sison, 2008). Mitchell (1960) concluded that samples of Al non-accumulating plants may be considered soil contaminated if soil: plant Al concentration ratio was less than 1,000:1. However, some metallophyte species may accumulate high concentrations of Al in their tissues without any signs of toxicity (Cherney and Robinson, 1983, Mossor-Pietraszewska, 2001), such as Melastoma malabathricum L. (Al accumulator) which stores more than 10,000 mg Al.kg-1 DW in its leaves and roots (Watanabe et al., 2008). Threshold concentrations between endogenous (absorbed by the plant) and exogenous (external contamination, such as by dust) Al have been identified as 70, 100 and 120 mg.kg-1 leaf DW for grass, clover, or other herbage, respectively (Metson et al., 1979), and Al concentration in leaves of most plant species as less than 200 mg.kg-1 DW (Mossor-Pietraszewska, 2001). Metson et al. (1979) found that there was a reasonable agreement between soil contamination estimated from Al and Fe values, although in some cases, values resulting from Fe were significantly higher than those from Al. Yasmin et al. (2014) concluded that Fe contamination on milled wheat seeds can be corrected by adjusting the Fe concentrations based on the concentration of Al in the grain.
Iron has also been suggested as an indicator of plant sample contamination (Metson et al., 1979, Mayland and Sneva, 1982). The common range of Fe in soils is between 0.1 and 10% and concentrations of easily soluble and exchangeable fractions of Fe are very low in comparison with the total Fe concentration, about 0.01 to 0.1% of the total Fe (Kabata-Pendias and Mukherjee, 2007). Mean Fe concentrations range from 43 to 376 mg.kg-1 DW in forage grasses and from 117 to 400 mg.kg-1 DW in clover (Kabata-Pendias and Mukherjee, 2007). Foliar Fe concentrations exceeding 500 mg.kg-1 DW can be considered toxic (Marschner, 2002, Fernando et al., 2009). Mayland and Sneva (1982) studied the effect of soil contamination on the mineral composition of forage fertilized with nitrogen (N) calculated from the dilution of soil Ti and Fe (assuming that the uncontaminated tissue contained 0 mg Ti.kg-1 DW and 80 mg Fe.kg-1 DW) and found that element concentrations in forage tissues calculated by Fe procedure are similar to those calculated by Ti procedure. However, the authors argued that N fertilizer increased Fe concentrations in the plant tissue (Mitchell, 1960). Iron is considered as a good indicator as it has a soil/plant ratio of 50,000/100 (i.e. 500) and its concentration in the plant is readily affected by soil contamination (Mitchell, 1960). However, as pH significantly affects forage Fe uptake, the effectiveness of Fe as indicator of soil contamination is reduced, especially for plants grown on soil with different pH (Metson et al., 1979, Mayland and Sneva, 1982). In addition, Metson et al. (1979) argued that because uptake of Fe by plants is variable, Fe is a less useful indicator of soil contamination
86 compared to Al and Ti. Similar to Ti, the concentration of Fe may also vary with soil particle size and extraction or solubilisation of soil and plant samples should use the same procedures to account for similar solubilities in the soil and dust fractions (Mayland and Sneva, 1982).
Scandium (Sc) has also been used as an indicator of soil contamination (Ferrari et al., 2006). Average concentrations of Sc in soils worldwide range between 1.5 and 16.6 mg.kg-1 (Kabata- Pendias and Mukherjee, 2007). Scandium cannot be taken up actively by plants (Shtangeeva et al., 2004) and the commonly reported range for Sc in various plants (leaves) is 0.002–0.25 mg.kg-1 DW (Kabata-Pendias and Mukherjee, 2007). Scandium was successfully used by Ferrari et al. (2006) to evaluate surface contamination of leaves from the Atlantic Forest of South America, and to correct and calculate intrinsic concentrations of cobalt (Co), caesium (Cs), potassium (K) and zinc (Zn) without contamination influence in leaves. However, for other elements such as Fe, this correction technique was found to be very poor (Ferrari et al., 2006).
Silicon (Si) is another indicator of soil contamination. However, its usefulness has been criticized because its uptake by plants is rather variable and it is a major constituent of graminaceous plants (Metson et al., 1979, Currie and Perry, 2007, Kabata-Pendias and Mukherjee, 2007). Kabata- Pendias and Mukherjee (2007) found that mean Si concentrations range between 0.3 and 1.2% DW in grasses and from 0.05-0.2% DW in leguminous plants. In sedges, nettles, horsetails and diatoms, Si concentrations can be up to >10% DW.
Soil contamination indicators have been ranked in order of sensitivity using soil/plant ratios (Table 3-2, and the most effective indicators are Ti>Cr>Al>Fe (Mitchell, 1960).
Table 3-2 Estimated total concentrations of selected metals in soil fine material and plant dry matter (mg.kg-1) and soil/plant ratios (Mitchell, 1960). Best soil contamination Soil/Plant Soil Plant indicators ratio Titanium 10,000 1 10,000 Chromium 200 0.1 2,000 Aluminium Not indicated Not indicated 1,000 Iron 50,000 100 500
Although residual contamination of our plant samples by soil was expected to be very low given that i) at the time of collection in the field, all leaf samples were collected well above the soil surface, and ii) all material including roots was thoroughly cleaned prior to laboratory analysis, all
87 samples were assessed for possible soil contamination using three sensitive indicators, Cr, Al, and Fe (Table 3-3). Titanium was not employed in this study as the elements Al, Ca, Cu, Fe, K, Mg, Mn, Na, P, S, and Zn were analysed via inductively coupled plasma optical emission spectroscopy (ICP-OES) and the elements Hg, Cr, Co, Ni, As, Se, Cd, Pb, Mo and Au were analysed via inductively coupled plasma mass spectrometry (ICP-MS). Therefore, the analytical sensitivities were not identical for all the elements (Mayland and Sneva, 1982). The ranges of lowest observable adverse effects concentrations (LOAECs) in Table 3-3 were derived from the species for which information was available. The threshold metal concentration in roots or leaves in Table 3-3 was set at 20 per cent above the highest LOAEC for the relevant attribute.
As described in Table 3-3, samples of Cr non-accumulating plants were considered contaminated using Cr as an indicator when Cr concentration in leaves exceeded 36 mg.kg-1 DW; soil/plant ratios of Cr were less than 2,000 (Mitchell, 1960); and shoot/root ratio of Cr exceeded 0.01 (Cary and Kubota, 1990, Sharma et al., 2005).
Samples of Al non-accumulating plants were considered contaminated using Al as an indicator when Al concentration exceeded 240 mg.kg-1 DW in leaves and 1,440 mg.kg-1 DW in roots (Stangoulis, 2010, Yasmin et al., 2014); soil/plant ratios of Al were less than 1,000 (Mitchell, 1960); and shoot/root ratio of Al exceeded 0.1 (Bouma et al., 1981).
Samples of Fe non-accumulating plants were considered contaminated using Fe as an indicator when Fe concentration exceeded 600 mg.kg-1 DW in leaves (i.e. 20% above the 500 mg.kg-1 DW foliar toxicity threshold of Marschner (Marschner, 2002, Fernando et al., 2009)) and 9,600 mg.kg-1 DW in roots; and soil/plant ratios of Fe were less than 500 (Mitchell, 1960).
A working rule was adopted whereby a plant sample was deemed to be contaminated by soil if the concentrations of the indicator elements (Al, Fe and Cr) all exceeded their respective concentration thresholds (Table 3-3). Based on this rule, 21 out of 318 samples of the 43 plant species that were analysed as part of this study were considered potentially contaminated by soil and data collected from these samples were excluded from subsequent analyses.
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Table 3-3 Lowest observable adverse effects concentrations (LOAECs) and associated thresholds used for monitoring possible soil contamination of plant samples. Values are not given for very sensitive or highly tolerant plant species. Concentrations of Cr, Al and Fe in soil refer to total concentrations of these metals in soil. “Plant” in soil/plant ratios refers to above ground plant parts. Indicators of Soil/plant Foliar Foliar Root Root Shoot/root Types of plants Plant parts contamination ratio LOAEC threshold LOAEC threshold concentration threshold (mg.kg-1 concentration (mg.kg-1 concentration ratio threshold DW) (mg.kg-1DW) DW) (mg.kg-1DW) Chromium <2,000a 5 - 30b >36 - - >0.01h Vegetable crops, pasture Leaves herbage Aluminium <1,000a 20c - 200d >240 >1,200f >1,440 >0.1f Grass, clover, pasture herbage, Leaves, rice rice, subterranean clover shoots, roots
Iron <500a 500e >600e >8,000g >9,600 - Grass, clover, pasture herbage, Leaves, roots bush beans, herbarium specimens of plants from the Myrtaceae, Proteaceae and Celastraceae families
Sources: a(Mitchell, 1960); b(Kabata-Pendias and Mukherjee, 2007); c(Wallace and Romney, 1977); d(Mossor-Pietraszewska, 2001); e(Marschner, 2002, Fernando et al., 2009); f(Bouma et al., 1981); g(Wallace et al., 1977); h(Cary and Kubota, 1990, Zayed et al., 1998, Sharma et al., 2005).
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3.3.5 Data analysis For field collected samples, data were presented as means and standard errors of three replicates for metal concentration in plants and soils. For determination of kinetic parameters, each data point represented each sample replicate (for field collected data) and between 2 to 27 replicates (for data obtained from the literature).
Unit conversions of the literature data were performed when necessary for uniformity of units throughout the study. Conversion of solution molar concentrations to soil dry weight concentration requires an assumption of soil water content. This fraction varies widely, from 1 to 30% (w/w), but during active plant growth in most soils it is between 10 and 20% (w/w) (Kirkham, 2014). A soil water content of 15% (w/w) was assumed for the period of active plant uptake (Fredlund and Xing, 1994) and was applied to field collected data and to the literature data (if required) when converting units of transition concentrations between high and low affinity metal transport systems.
Arsenic, Au, Cr, Co, Cd, Hg, Mo, Pb and Se are normally found only at low concentrations in the environment, so their concentrations in plant parts were expected to be low. These elements are normally measured by ICP-MS, which has a higher analytical sensitivity than ICP-OES. However, plant-available metals in soil samples were extracted by CaCl2, and its high concentration in the extractant precluded the use of the more sensitive ICP-MS. The dependence on the low analytical sensitivity of ICP-OES meant that As, Au, Cr, Co, Cd, Hg, Mo, Pb and Se were omitted from further analysis, which focused only on Al, Cu, Fe, Mn, Ni and Zn.
Uptake parameters determined from field collections in this study assume that they are made from plant materials of uniform age. Greater uniformity may be achieved by classifying material into structural (shoots, leaves, roots) and age cohorts (living or senescent leaves) where this could be done reliably and where there was sufficient material to provide samples for analysis. Uptake parameters were determined independently for each cohort. Material from these different cohorts may be expected to have different final metal concentrations and different apparent uptake parameters if they are assessed against a common uptake period.
For grass species collected in central Queensland, it can be assumed that live leaves were produced during a 3- to 4-month summer growing season, and that senescent leaves had maintained active metal uptake for a similar period of time. As a result, it would be expected that when metal uptake ceased on leaf death, the concentrations in leaves would remain approximately constant or could possibly decline due to leaching or volatilisation.
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For dicotyledonous (evergreen) species, many that occur in central Queensland retain leaves for approximately one year, with one-year-old leaves being shed soon after the initiation of a new season’s growth (authors’ observations). Low soil water contents during the dry season are also associated with markedly reduced transpiration and commensurately limited nutrient and metal uptake. Therefore, dicotyledonous species collected after the end of the major growing season would be expected to have foliar metal concentrations that reflected their growing season conditions.
For these reasons, it is argued that foliar metal concentrations in field-grown plants reflect the plant available metal concentrations in the soil, a parameter that is determined at a nominal soil water content of 15% (w/w), which represents growing season rather than dry season conditions.
3.4 Results 3.4.1 Conceptual framework for plant metal uptake from soils Based on a review of the literature, we have developed, for the first time, a functional conceptual framework of plant metal uptake in relation to plant available metal concentration in the soil (Figure 3-1). This framework applies to all plant parts, and not only leaves.
Figure 3-1 Conceptual model of changes in metal uptake into plants, integrated over time, as a function of low and high plant available metal concentrations in the substrate. Experimental (EXP) plants grown in controlled laboratory conditions in solution are either deprived (grown
91 without metal), initiating a constitutive high-affinity transport system (CHATS) at low experimental metal concentrations and both constitutive high- and low-affinity transport systems (CLATS+CHATS) at high experimental metal concentrations, (brown lower curve), or metal-supplied (grown in the presence of metal), initiating an inducible high-affinity transport system (IHATS) at low experimental metal concentrations and both inducible high- and low-affinity transport systems (ILATS+IHATS) at high experimental metal concentrations, (blue middle curve). Uptake attributed to the CHATS or IHATS follows
Michaelis-Menten kinetics with parameters Vmax and Km (the saturable component). When substrate metal concentrations reach the transition concentration (shown by the vertical line above the asterisk in the figure), low-affinity transport systems (CLATS and ILATS) whereby metal uptake increases linearly with increasing substrate concentration (the linear component), dominate total metal uptake (CHATS+CLATS and IHATS+ILATS). The linear portions of the uptake curves are identified by the slope and intercept (crosses at zero- concentration) for each transport system. For plants collected from the field, uptake will be the combined result of both constitutive and inducible uptake (FIELD HATS at substrate plant available metal concentrations below the transition concentration and FIELD HATS+LATS at greater substrate plant available metal concentrations, red upper curves). At very high external metal concentrations, mechanisms of plant metal resistance or tolerance break down and unrestricted transport occurs causing plant death.
The framework combines saturable Michaelis-Menten kinetics (HATS) and linear metal uptake (LATS). Two constitutive systems operate in non-induced plants: a constitutive high-affinity transport system (CHATS) at low external concentration of metals, and a constitutive low-affinity transport system (CLATS) at high external metal concentration (≥ transition concentration between transport systems). When plants are induced by exposure to certain concentrations of metals, plant uptake increases over the whole range of concentrations. An inducible high-affinity transport system (IHATS) approaches Michaelis-Menten kinetics up to concentrations below the transition concentration between transport systems. At greater concentrations, a putative inducible low- affinity transport system (ILATS) is activated and dominates metal uptake into plants.
We propose that five parameters can be used to describe metal uptake in plant parts grown in solution in controlled laboratory conditions. Depending on the uptake components, maximum uptake rate (Vmax) of metal and affinity of transporters (Km) are calculated for both high affinity components, while the slopes and intercepts are calculated for the linear components. Transition concentration between transport systems occurs where constitutive or inducible low affinity
92 transport systems dominate the uptake and show linear pattern at high external metal concentrations
(Figure 3-1). Vmax determines the extent of metal accumulation in plant parts while Km is the affinity of a plant part for the metal. Thus, the lower the Km of a plant part, the higher its affinity, resulting in preferential accumulation of the metal in that plant part. The linear slopes of the low affinity transport systems reflect the extent of accumulation in each plant part over a given time period at high external metal concentrations. At extreme soil metal concentrations, plant metal uptake control mechanisms break down and unrestricted transport occurs, leading to metal toxicity symptoms and eventually plant death.
We propose further that for field collected plants, time-integrated metal uptake rates can be expressed as concentrations, so that Michaelis-Menten and linear uptake kinetics can be used to describe the relationships between metal concentrations in field-collected plant parts and available metal concentrations in the soil. For field-grown material, there is no means to separate the induced and non-induced uptake, so the field responses reflect the combination of the two processes.
We suggest new terms for the classification of metallophytes according to their accumulation characteristics. Plants that grow and survive in soils that are toxic to other plants without displaying any toxicity symptoms are called metal “tolerators”. These plants have evolved various mechanisms of tolerance that regulate and detoxify metals in different cellular compartments in shoots/leaves and/or in roots and/or that limit transport of metals to sensitive plant parts. There are three types of tolerators: i) shoot/leaf tolerators, ii) root tolerators, and iii) shoot/leaf-root tolerators. A shoot or leaf tolerator is defined as a plant species that primarily accumulates metals in its leaves or shoots compared to the roots. In contrast, a root tolerator is a plant species that limits metal transport into aboveground tissues and stores metals mainly in its roots compared to the shoots/leaves. A plant species is defined as a shoot/leaf and root tolerator if accumulation of metals occurs in both its aerial and belowground parts.
Depending on the plant available concentrations of metals in the soil, tolerator responses to metals can be described best by either Vmax and Km for soil concentrations below the transition concentration between transport systems (HATS), or by the slope and intercept for soil available metal concentrations greater than the transition concentration between the two transport systems (LATS) (Figure 3-1).
Table 3-4 describes element (plant essential and non-essential) transporter kinetic parameter values in plants collected from the literature, in relation to their metal tolerance characteristics (for metal elements). Most studies were found to relate to HATS and studies on LATS were scarce (Table 3-
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4). Studies on uptake kinetics of Zn by the Zn-hyperaccumulator species Thlaspi caerulescens and the non-accumulator species T. arvense showed that influx of Zn into roots of the former species was higher than the latter species after 20 minutes of exposure to Zn in solution at up to the transition concentration between transport system of 100 μM, with Vmax of T. caerulescens being 4.5-fold greater than T. arvense (Lasat et al., 1996) (Table 3-4). Other studies showed influx of As(V) into whole plants of the As-hyperaccumulator fern Pteris vittata and the non-accumulator fern Nephrolepis exaltata (Table 3-4). Both species had similar Vmax values either without or with
30 μM phosphate, but the Km of the non-accumulator species N. exaltata were 9 and 3 times greater, respectively, than that of the hyperaccumulator species P. vittata (Poynton et al., 2004).
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Table 3-4 Transporter kinetic parameters of essential and non-essential elements in plants, extracted from the literature. Values are given as the mean ± standard errors (when available from the articles). The number of replicates varied from 2 to 27. Element Plant species Plant Treatment* HATS parameters† Transition LATS parameters†† 2 2 and element Vmax Km R concn** Slope Intercept R reference tolerance (µM) (µmole g (µmole g root (µmol g root characteris root DW-1 DW-1 h-1) DW-1 h-1) tics h-1µM -1) Essential Macronutrient Mg2+ Oryza sativa n.a Mg-deficient 2.16 (IH) 70 (IH) n.d 100a n.d n.d n.d L. cv. Nipponbarea n.a Mg-sufficient 1.68 (IH) 257 n.d 100a n.d n.d n.d (IH) + + c; d NH4 Camellia n.a - NH4 3.27 (CH) 60 n.d 100 n.d n.d n.d sinensis L.b (CH) Oryza sativa n.a 4-week-old 219.6 (IH) 75 (IH) n.d 100c; d n.d n.d n.d var IR-36e plant n.a 9-week-old 61.2 (IH) 103 n.d 100c; d n.d n.d n.d plant (IH) Picea glauca n.a N-deprived 18.6 (CH) 19.79 0.93 100c; d n.d n.d n.d (Moench) (CH) (CH) Vossf (Linew eaver- Burk) - - g; h NO3 Brassica n.a - NO3 26.3 (CH) 15.9 n.d 135-150 0.007 19.6 n.d napus L. cv. (CH) (CL+CH) (CL+CH) Capitalg - g; h n.a + NO3 135 (IH) 85 (IH) n.d 135-150 0.01 124 (IL+IH) n.d (IL+IH) - g; h Camellia n.a - NO3 2.358 (CH) 160 n.d 135-150 n.d n.d n.d sinensis L.b (CH)
95
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µmol g (µM) (µmole g (µmole g root DW- root DW-1 root DW-1 1 h-1) h-1µM -1) h-1) Essential Macronutrient - j g; h NO3 Citrus reshni n.a Troyer 2.7±0.2 282±6 n.d 135-150 2.38 (CL) n.d n.d citrange (CH) (CH) n.a Cleopatra 2.5±0.1 281±8 n.d 135-150g; h 1.73 (CL) n.d n.d mandarin (CH) (CH) Hordeum n.a IND 100 μM 94.10 79 n.d 135-150g; h n.d n.d n.d - vulgare L. cv NO3 (IH) (IH) Klondikek (1 day) n.a IND 100 μM 36.30 45 n.d 135-150g; h n.d n.d n.d - NO3 (IH) (IH) (4 days) n.a IND 10000 16.90 30 n.d 135-150g; h n.d n.d n.d - μM NO3 (IH) (IH) (4 days) n.a UNIN 3.44 20 n.d 135-150g; h n.d n.d n.d (CH) (CH) n.a IND 100 μM n.d n.d n.d 135-15-g; h 0.0057 n.d n.d - NO3 (IL) (1 day) n.a UNIN n.d n.d n.d 135-150g; h 0.0064 n.d n.d (CL) Oryza sativa n.a 4-week-old 158.4 75 n.d 135-150g; h n.d n.d n.d var IR-36e plant (IH) (IH)
96
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Macronutrient - g; h NO3 Picea glauca n.a UNIN 1.1 (CH) 13.6 0.91 135-150 n.d n.d n.d (Moench) (CH) (CH) Voss.h (Linew eaver- Burk) n.a IND 3.2 (IH) 16.83 0.97 135-150g; h n.d n.d n.d (saturabl (IH) (IH) e comp. (satur (Linew 1) able eaver- comp. Burk) 1) n.a IND 8.2 (IH) 153.1 0.9 135-150g; h n.d n.d n.d (saturabl 7 (IH) (IH) e comp. (satur (Linew 2) able eaver- comp. Burk) 2) n.a UNIN n.d n.d n.d 135-150g; h 0.000001 3.3 (CL) 0.97 (CL) 5 (CL) (linear regression, student’s t- tests)
97
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Macronutrient - - g; h NO3 Zea mays L. n.a UNIN –NO3 50.2±4.3 86±2 n.d 135-150 n.d n.d n.d cv. Cecilia; (CH) (CH) - g; h Pioneer Hi- n.a IND +NO3 123.3±23 30±2 n.d 135-150 n.d n.d n.d Bred Italia (IH) (IH) - g; h Srl, Pieve n.a UNIN –NO3 , 62.9±17. 74±10 n.d 135-150 n.d n.d n.d Delmona, 10 min Cd2+ 9 (CH) (CH) l - g; h CR, Italy n.a IND +NO3 , 94.8±12. 79±16 n.d 135-150 n.d n.d n.d 10 min Cd2+ 2 (IH) (IH) n.a UNIN – 55.1±16. 89±2 n.d 135-150g; h n.d n.d n.d - NO3 , 12 h 3 (CH) (CH) Cd2+ - g; h n.a IND +NO3 , 78.7±7.5 94±0. n.d 135-150 n.d n.d n.d 12 h Cd2+ (IH) 4 (IH)
98
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Micronutrient Cu2+ Hordeum No inf 111.1 23400 n.d 100m; n n.d n.d n.d vulgare L., (CH) (CH) cv. Herta (phase 1)m Hordeum No inf 5.48 13.2 n.d 100m; n n.d n.d n.d vulgare L., (CH) (CH) cv. Herta (phase 2)m Mimulus Cu-tolerant 2.58 (IH) 7.82 0.902 100m; n n.d n.d n.d guttatus (IH) Fischer ex Cu-non tolerant 5.05 (IH) 21.36 0.902 100m; n n.d n.d n.d DCp (IH) Saccharun No inf 5.37 14.5 n.d 100m; n n.d n.d n.d officinarum (CH) (CH) L., var. H53- 263n
99
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Micronutrient Fe2+ Zea mays L., Fe-efficient UNIN -Fe 11 (CH) 9.6 n.d 30q n.d n.d n.d cv Aliceq (CH) Fe-efficient IND 8.6 (IH) 10.9 n.d 30q n.d n.d n.d +FeEDTA (IH) Zea mays L., Fe-inefficient UNIN -Fe 1.2 (CH) 10.7 n.d 30q n.d n.d n.d ysl mutantq (CH) Fe-inefficient IND 0.84 (IH) 8.2 n.d 30q n.d n.d n.d +FeEDTA (IH)
100
Table 3-4. contd. Element Plant species Plant element Treatment HATS parameters† Transitio LATS parameters†† 2 2 and tolerance * Vmax Km R n concn** Slope Intercept R reference characteristics (µmol g (µM) (µmole g (µmole g root DW-1 root DW-1 root DW-1 h-1) h-1µM -1) h-1) Essential Micronutrient Mn2+ Citrus No inf 0.06±0.03 5.3±2.4 n.d 0.13s n.d n.d n.d aurantium L. (IH) (IH) (phase 1)r Citrus No inf 0.19±0.03 7.2±3.3 n.d 0.13s n.d n.d n.d aurantium L. (IH) (IH) (phase 2)r Citrus No inf 1.63±0.04 64±4 n.d 0.13s n.d n.d n.d aurantium L. (IH) (IH) (phase 3)r Hordeum Tolerant to low Mn- 0.0054±0. 0.0054 0.980 0.13s n.d n.d n.d vulgares concentration of efficient 0010 (IH) ±0.001 (IH) Mn 8 (IH) (Regressi on analysis via SAS program) Not tolerant to Mn- 0.0014±0. 0.0027 0.998 0.13s n.d n.d n.d low inefficient 0004 (IH) ±0.001 (IH) concentration of 6 (IH) Regressi Mn on analysis via SAS program)
101
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Micronutrient Mn2+ Hordeum No inf 41.7 845 n.d 0.13s n.d n.d n.d vulgare (CH) (CH) (phase 1)m Hordeum No inf 5.17 5.68 n.d 0.13s n.d n.d n.d vulgare (CH) (CH) (phase 2)m Oryza sativa No inf 5.3 (CH) 630 n.d 0.13s n.d n.d n.d L. cv. I.R. 8t (CH) Saccharun No inf 5.35 16.1 n.d 0.13s n.d n.d n.d officinarum (CH) (CH) L., var. H53- 263n Ni2+ Avena sativa No inf 2.72 12 0.98 10u n.d n.d n.d L. cv. (CH) (CH) (CH) Victoryu (Haines -Woolf plots)
102
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Micronutrient Zn2+ Hordeum No inf 20.4 420 n.d 20v n.d n.d n.d vulgare (CH) (CH) (phase 1)m Hordeum No inf 5.71 16 n.d 20v n.d n.d n.d vulgare (CH) (CH) (phase 2)m Oryzae sativa Zn-efficient 29 (CH) 6 n.d 20v n.d n.d n.d L.w (CH) Zn-inefficient 57.4 13 n.d 20v n.d n.d n.d (CH) (CH) Saccharun No inf 5.88 11.1 n.d 20v n.d n.d n.d officinarum (CH) (CH) L., var. H53- 263n Thlaspi Zn 2.7 (IH) 8 (IH) n.d 20v 0.0357 n.d n.d caerulescensv hyperaccumulat (IL) or Thlaspi Zn non- 0.6 (IH) 6 (IH) n.d 20v 0.0315 n.d n.d arvensev accumulator (IL)
103
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Essential Micronutrient Zn2+ Triticum Low-Zn tolerant Zn-deficient 0.109±0. 0.6±0. n.d 20v n.d n.d n.d aestivum L.y (Zn-efficient) 015 (IH) 6 (IH) Low-Zn tolerant Zn-sufficient 0.09±0.0 1.2±0. n.d 20v n.d n.d n.d (Zn-efficient) 06 (IH) 5 (IH) Low-Zn Zn-deficient 0.309±0. 2.3±1 n.d 20v n.d n.d n.d intolerant (Zn- 022 (IH) (IH) inefficient) Low-Zn Zn-sufficient 0.097±0. 0.7±0. n.d 20v n.d n.d n.d intolerant (Zn- 01 (IH) 5 (IH) inefficient)
Non-essential
Al3+ No inf 25z As (Arsenate) Holcus As tolerant 0.672±0. 565±3 n.d 100aa; ab n.d n.d n.d lanatus L.aa 195 (CH) 34 (CH) As (Arsenate) As non-tolerant 1.578±0. 24±11 n.d 100aa; ab n.d n.d n.d 169 (CH) (CH)
104
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
As (Arsenate) Holcus As tolerant Gen a 2.018±0. 25±14 n.d 100aa; ab n.d n.d n.d lanatus L.ac 439 (CH) (CH) As (Arsenate) As tolerant Gen b 3.518±0. 42±20 n.d 100aa; ab n.d n.d n.d 706 (CH) (CH) As (Arsenate) As tolerant Gen c 3.165±0. 40±20 n.d 100aa; ab n.d n.d n.d 683 (CH) (CH) As (Arsenate) As tolerant Gen d 3.061±0. 18±8 n.d 100aa; ab n.d n.d n.d 453 (CH) (CH) As (Arsenate) As tolerant Gen e 2.463±0. 30±12 n.d 100aa; ab n.d n.d n.d 374 (CH) (CH) As (Arsenate) As tolerant Gen f 2.864±0. 25±10 n.d 100aa; ab n.d n.d n.d 442 (CH) (CH) As (Arsenate) As non-tolerant Gen g 2.99±0.1 22±7 n.d 100aa; ab n.d n.d n.d 76 (CH) (CH) As (Arsenate) As non-tolerant Gen h 4.133±0. 35±11 n.d 100aa; ab n.d n.d n.d 549 (CH) (CH) As (Arsenate) As non-tolerant Gen i 4.978±0. 45±17 n.d 100aa; ab n.d n.d n.d 844 (CH) (CH) As (Arsenate) As non-tolerant Gen j 4.942±0. 36±15 n.d 100aa; ab n.d n.d n.d 879 (CH) (CH) As (Arsenate) As non-tolerant Gen k 5.4±1.15 38±19 n.d 100aa; ab n.d n.d n.d (CH) (CH) As (Arsenate) As non-tolerant Gen l 4.53±1.0 49±23 n.d 100aa; ab n.d n.d n.d 11 (CH) (CH)
105
Table 3-4. contd. Element Plant species Plant element Treatme HATS parameters† Transit LATS parameters†† 2 2 and tolerance nt* Vmax Km R ion Slope Intercept R reference characteristics concn* * (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW-1 h- 1 h-1µM - h-1) 1) 1) Non-essential
As (Arsenate) Oryza sativa No inf 5.71±1.06 64.9±18.8 0.992 100aa; ab 0.01892 15.53 (CL) 0.990 L. (cultivar (CH) (CH) (CH) (CL) (CL) Guangyinzha (Anova, (Anova, )ab SPSS SPSS package) package) As (Arsenite) No inf 4.49±0.65 39.7±10.6 0.990 100aa; ab 0.01314 8.99 (CL) 0.946 (CH) (CH) (CH) (CL) (CL) (Anova, (Anova, SPSS SPSS package) package) As (DMA) No inf 2.95±0.83 72.2±30.7 0.985 100aa; ab 0.00415 2.94 (CL) 0.984 (CH) (CH) (CH) (CL) (CL) (Anova, (Anova, SPSS SPSS package) package) As (MMA) No inf 8.28±0.99 76.4±13.6 0.998 100aa; ab 0.02212 1.94 (CL) 0.988 (CH) (CH) (CH) (CL) (CL) (Anova, (Anova, SPSS SPSS package) package)
106
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
As (Arsenate) Oryza sativa No inf 24.98±4. 89.4± 0.996 100aa; ab 0.00985 11.76 (CL) 0.977 L. (cultivar 72 (CH) 22.5 (CH) (CL) (CL) Handao (CH) (Anova (Anova 502)ab , SPSS , SPSS packag packag e) e) As (Arsenite) No inf 6.78±1.5 71.4± 0.990 100aa; ab 0.01509 1.28 (CL) 0.963 3 (CH) 24.5 (CH) (CL) (CL) (CH) (Anova (Anova , SPSS , SPPS packag packag e) e) As (DMA) No inf 1.7±0.21 17.9± 0.977 100aa; ab 0.00182 1.5 (CL) 0.886 (CH) 5.5 (CH) (CL) (CL) (CH) (Anova (Anova , SPSS , SPSS packag packag e) e) As (MMA) No inf 4.72±1.1 40.6± 0.974 100aa; ab 0.01328 4.12 (CL) 0.973 (CH) 17.2 (CH) (CL) (CL) (CH) (Anova (Anova , SPSS , SPSS packag packag e) e)
107
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
As (Arsenate) Lupinus No inf +P (1 week) 1.25±0.1 37.7± 0.9920 100aa; ab n.d n.d n.d albusad 6 (CH) 8.8 (CH) (CH) (Marqu ardt– Levenb erg algorith m) As (Arsenate) Lupinus No inf -P (1 week) 1.58±0.1 26.5± 0.9961 100aa; ab n.d n.d n.d albusad 0 (CH) 3.7 (CH_ (CH) Marqua rdt– (Leven berg algorith m)
108
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
As (Arsenate) Lupinus No inf +P (2 week) 1.81±0.6 67.4± 0.9761 100aa; ab n.d n.d n.d albusad 4 (CH) 37.2 (CH) (CH) (Marqu ardt– Levenb erg algorith m) As (Arsenate) Lupinus No inf -P (2 week) 1.62±0.0 16.2± 0.9998 100aa; ab n.d n.d n.d albusad 1 (CH) 0.5 (CH) (CH) (Marqu ardt– Levenb erg algorith m)
109
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
As (Arsenate) Pteris vittata As UNIN -P 0.12±0.0 1.1±0. 0.85 100aa; ab n.d n.d n.d L.ae hyperaccumulat 18 (CH) 3 (CH) or (CH) (Hanes- Woolf linear plots) As (Arsenate) Nephrolepis As non- UNIN -P 0.144±0. 9.9±2. 0.70 100aa; ab n.d n.d n.d exaltata L.ae accumulator 02 (CH) 0 (CH) (CH) (Hanes- Woolf linear plots) As (Arsenate) Pteris vittata As IND 0.095±0. 6.8±1. 0.83 100aa; ab n.d n.d n.d L.ae hyperaccumulat +30 μM P 008 (CH) 9 (CH) or (CH) (Hanes- Woolf linear plots) As (Arsenate) Nephrolepis As non- IND 0.097±0. 19.9± 0.73 100aa; ab n.d n.d n.d exaltata L.ae accumulator +30 μM P 026 (CH) 3.3 (CH) (CH) (Hanes- Woolf linear plots)
110
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-essential
Cd2+ Pisum No inf IND +Fe 0.34±0.0 0.6±0. n.d 10af 0.053±0. n.d 0.999 sativum L. cv 8 (CH) 9 0013 (CL) Sparkle (Fe- (CH) (CL) (Least sufficient)af square proce dure, sigma Plot) Pisum No inf UNIN –Fe 2.36±0.2 1.5±0. n.d 10af 0.067±0. n.d 0.999 sativum L. cv 5 (CH) 6 004 (CL) (CL) Sparkle (Fe- (CH) (Least deficient)af square proce dure, sigma Plot) Thlaspi Cd 1.432±0. 0.45± 0.91 10af n.d n.d n.d caerulescens hyperaccumulat 066 (CH) 0.07 (CH) (Ganges or (CH) (ANOV ecotype)ag A)
111
Table 3-4. contd. Element Plant species Plant element Treatment* HATS parameters† Transition LATS parameters†† 2 2 and tolerance Vmax Km R concn** Slope Intercept R reference characteristics (µM) (µmole g (µmole g (µmol g root DW- root DW-1 root DW- 1 h-1µM - h-1) 1 h-1) 1) Non-Essential Co2+ No inf Cr(III) Leerdia No inf Fe deficiency 20.7 95.1 0.97 150ah n.d n.d n.d hexandra (CH) (CH) (CH) Swartzah (Linew eaver- Burk) No inf Fe(III) supply 15.2 252.3 0.97 150ah n.d n.d n.d (CH) (CH) (CH) (Linew eaver- Burk) Pb2+ No inf
* Treatment abbreviations: Gen, genotype; IND, induced; UNIN, uninduced; ** Transition concentration between metal transport systems. † High-affinity transport system parameters (CH, constitutive; IH, inducible) †† Low-affinity transport system parameters (CL, constitutive; IL, inducible) Abbreviations: n.a., not applicable; n.d., not determined; no inf, no information found; DMA, Dimethylarsinic acid; MMA, monomethylarsonic acid. References: a(Tanoi et al., 2014); b(Yang et al., 2013b); c(Wang et al., 1993); d(BassiriRad et al., 1999); e(Youngdahl et al., 1982); f(Kronzucker et al., 1996); g(Faure-Rabasse et al., 2002); h(Kronzucker et al., 1995); j(Cerezo et al., 1997); k(Siddiqi et al., 1990); l(Rizzardo et al., 2012); m(Bowen, 1981);
112 n(Bowen, 1969); p(Strange and Macnair, 1991); q(Wiren et al., 1995); r(Hassan, 1977); s(Pedas et al., 2005); t(Ramani and Kannan, 1975); u(Aschmann and Zasoski, 1987); v(Lasat et al., 1996); w(Bowen, 1986); y(Hacisalihoglu et al., 2001); z(Guerrier, 1978); aa(Meharg and Macnair, 1992a); ab(Li et al., 2011b); ac(Meharg and Macnair, 1992b); ad(Esteban et al., 2003); ae(Poynton et al., 2004); af(Cohen et al., 1998); ag(Zhao et al., 2002b); ah(Liu et al., 2011)
113
3.4.2 Conceptual framework validation using field survey of Queensland native plant species
In this section, we focus on six metals (Al, Cu, Fe, Mn, Ni and Zn) that are generally found to be abundant in metal-enriched areas. There were poor correlations between total metal concentration and plant available metal concentration in soils (Table 3-5). For example, total concentration of Mn in the serpentine area of Rockhampton (Site 1/2) and the Cracow mine (Site WTP237) were 2453±59 and 1659±17 mg total Mn.kg-1 soil DW but availabilities of the metal in those sites were 11.66±0.10 and 48.64±0.63 mg plant available Mn.kg-1 soil DW, respectively, or only 0.48±0.01 and 2.93±0.05% of the total, respectively. In contrast, total Mn concentration in Mt. Morgan mine (Site Mor-1) was 208±6 mg total Mn.kg-1 soil DW but its availability was high (12.35±0.27 mg plant available Mn.kg-1 soil DW) or 5.96±0.21% of the total. Similarly, total soil Ni concentrations in two different sites of serpentine areas of Rockhampton were 2030±23 and 5732±527 mg total Ni.kg-1 soil DW for Sites 1/2, and 2/4, respectively, but plant available Ni concentrations were 31.49±0.04 mg Ni.kg-1 soil DW (1.55±0.02%) and 7.73±0.19 mg Ni.kg-1 soil DW (0.14±0.01%), respectively (Table 3-5). Overall, only small fractions of total metals in soil samples collected from 43 sites in serpentine areas of Rockhampton and Cracow and Mt. Morgan mines in central Queensland, Australia, were available for plant uptake.
114
Table 3-5 Metal concentrations in soil at each plant species collection site. Values are given as the mean ± standard errors (n=3). Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil metal (mg.kg-1 DW) † (mg.kg-1 DW) available Al 1/7 Myrtaceae Corymbia xanthope 0.1011a 1.98±0.38 11845±403 0.02±0.00 2/1 0.1011a 0.22±0.18 15408±3088 0.00±0.00 5/15 0.1011a 2.24±0.19 20645±1599 0.01±0.00 1/4 Myrtaceae Eucalyptus fibrosa 0.1011a 1.34±0.2 11579±96 0.01±0.00 1/5 0.1011a 2.46±0.07 13166±177 0.02±0.00 5/18 0.1011a 0.38±0.08 20920±6931 0.00±0.00 8/11 0.1011a 3.68±0.41 44501±1235 0.01±0.00 Mor-1 Myrtaceae Eucalyptus 0.1011a 275.2±2.28 46786±3260 0.59±0.05 tereticornis 4/3 0.1011a 0.86±0.29 46385±216 0.00±0.00 7/2 0.1011a 0.58±0.04 45595±1317 0.00±0.00 4/9 Poaceae Aristida vagans 0.1011a 8.62±1.39 40410±1523 0.02±0.00 5/10 0.1011a 0.85±0.22 29718±2679 0.00±0.00 4/7 Poaceae Sorghum nitidum 0.1011a 1.19±0.23 47910±334 0.00±0.00 forma aristatum 6/3 0.1011a 4.87±1.32 26216±2188 0.02±0.00 1/11 Poaceae Themeda triandra 0.1011a 1.69±0.13 12420±161 0.01±0.00 8/6 0.1011a 3.90±0.62 37005±416 0.01±0.00
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Table 3-5 contd. Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil me tal (mg.kg-1 DW) † (mg.kg-1 DW) available Cu 4/9 Poaceae Aristida vagans 0.95319b, c 0.19±0.03 13±0 1.5±0.25 5/10 0.95319b, c 0.04±0.01 36±1 0.11±0.02 1/4 Myrtaceae Eucalyptus fibrosa 0.95319b, c 0.16±0.02 24±0 0.67±0.07 5/18 0.95319b, c 0.07±0.01 46±1 0.16±0.01 4/9 Poaceae Aristida vagans 0.95319b, c 0.19±0.03 13±0 1.5±0.25 5/10 0.95319b, c 0.04±0.01 36±1 0.11±0.02 WTP239 Poaceae Melinis repens 0.95319b, c 1±0.02 32±1 3.14±0.1 WTP243 0.95319b, c 3.2±0.03 128±0 2.51±0.02 1/11 Poaceae Themeda triandra 0.95319b, c 0.05±0.01 27±6 0.2±0.05 8/6 0.95319b, c 0.25±0.01 47±0 0.53±0.03 Fe WTP237 Mimosaceae Acacia salicina 0.2513d 1.31±0.06 45238±1159 0.00±0.00 Mor-5 0.2513d 6.88±1.83 49111±829 0.01±0.00 Mor-1 Myrtaceae Eucalyptus 0.2513d 12.83±1.09 65086±801 0.02±0.00 tereticornis 4/3 0.2513d 0.95±0.21 23665±103 0.00±0.00 7/2 0.2513d 0.68±0.08 28048±344 0.00±0.00
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Table 3-5 contd. Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil metal (mg.kg-1 DW) † (mg.kg-1 DW) available Fe 2/10 Poaceae Panicum mitchellii 0.2513d 6.81±1.71 129498±2933 0.01±0.00 8/12 0.2513d 0.40±0.09 66675±2659 0.00±0.00 4/7 Poaceae Sorghum nitidum 0.2513d 1.49±0.24 33611±263 0.00±0.00 forma aristatum 6/3 0.2513d 8.10±0.14 45133±297 0.02±0.00 Mn 2/7 Cyperaceae Scleria mackaviensis 0.0010e 0.21±0.03 1570±6 0.01±0.00 4/8 0.0010e 1.77±0.12 1147±4 0.15±0.01 5/12 0.0010e 2.65±0.28 1279±3 0.21±0.02 8/8 0.0010e 13.73±0.77 2373±767 0.84±0.41 1/2 Mimosaceae Acacia leptostachya 0.0010e 11.66±0.10 2453±59 0.48±0.01 1/3 0.0010e 12.56±0.40 4807±85 0.26±0.01 2/4 0.0010e 2.96±0.09 2146±10 0.14±0.00 5/9 0.0010e 10.17±0.25 1688±12 0.6±0.02 WTP237 Mimosaceae Acacia salicina 0.0010e 48.64±0.63 1659±17 2.93±0.05 Mor-5 0.0010e 2.26±0.34 661±27 0.34±0.04 1/12 Poaceae Heteropogon 0.0010e 8.65±0.17 3248±49 0.27±0.01 contortus 4/5 0.0010e 0.94±0.16 904±13 0.1±0.02
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Table 3-5 contd. Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil metal (mg.kg-1 DW) † (mg.kg-1 DW) available e Mn 2/10 Poaceae Panicum mitchellii 0.0010 0.42±0.04 1707±26 0.02±0.00 8/12 0.0010e 7.42±0.09 2724±131 0.27±0.01 4/7 Poaceae Sorghum nitidum 0.0010e 1.77±0.12 1147±4 0.15±0.01 forma aristatum 6/3 0.0010e 0.67±0.15 1440±31 0.05±0.01 3/1 Thymelaeaceae Pimelea 0.0010e 16.69±0.08 3291±25 0.51±0.00 leptospermoides 5/6 0.0010e 5.3±0.02 1844±25 0.29±0.00 8/1 0.0010e 1.34±0.04 3030±63 0.04±0.00 Ni 2/7 Cyperaceae Scleria mackaviensis 0.08804f 0.54±0.09 4417±132 0.01±0.00 4/8 0.08804f 0.33±0.05 655±6 0.05±0.01 5/12 0.08804f 6.17±0.42 2799±58 0.22±0.01 8/8 0.08804f 0.49±0.01 744±238 0.09±0.04 1/2 Mimosaceae Acacia leptostachya 0.08804f 31.49±0.04 2030±23 1.55±0.02 1/3 0.08804f 17.43±0.35 1751±68 1.00±0.04 2/4 0.08804f 7.73±0.19 5732±527 0.14±0.01 5/9 0.08804f 3.38±0.06 3040±65 0.11±0.00 8/9 0.08804f 0.37±0.03 1006±21 0.04±0.00
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Table 3-5 contd. Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil metal (mg.kg-1 DW) † (mg.kg-1 DW) available Ni 1/6 Myrtaceae Corymbia xanthope 0.08804f 22.47±0.17 1976±93 1.14±0.05 2/1 0.08804f 2.28±0.13 3786±268 0.06±0.00 5/15 0.08804f 5.38±0.23 2940±124 0.18±0.01 1/4 Myrtaceae Eucalyptus fibrosa 0.08804f 7.75±0.07 1816±22 0.43±0.01 1/5 0.08804f 21.93±0.18 2006±54 1.1±0.04 5/18 0.08804f 4.55±0.07 6193±297 0.07±0.00 8/11 0.08804f 0.86±0.01 851±15 0.10±0.00 2/2 0.08804f 6.68±1.57 3472±43 0.19±0.05 1/10 Poaceae Cymbopogon refractus 0.08804f 21.94±0.24 2444±99 0.9±0.04 8/4 0.08804f 0.36±0.07 1010±16 0.04±0.01 1/12 Poaceae Heteropogon contortus 0.08804f 6.02±0.04 2726±254 0.23±0.02 4/5 0.08804f 0.42±0.26 418±6 0.1±0.06 8/7 0.08804f 0.5±0.03 831±40 0.06±0.01 3/1 Thymelaeaceae Pimelea 0.08804f 29.49±0.48 6516±134 0.45±0.01 leptospermoides 5/6 0.08804f 4.53±0.1 2792±224 0.16±0.01 8/1 0.08804f 0.06±0.02 1050±18 0.01±0.00
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Table 3-5 contd. Transition concn* Plant available Total metal in % of total Metal Site Family Plant species (mg.kg-1 DW) metal in soil soil metal (mg.kg-1 DW) † (mg.kg-1 DW) available Zn WTP237 Mimosaceae Acacia salicina 0.19614g 8.64±0.11 1823±45 0.47±0.01 Mor-5 0.19614g 0.17±0.06 206±3 0.08±0.03 1/4 Myrtaceae Eucalyptus fibrosa 0.19614g 0.05±0.00 72±1 0.07±0.01 1/5 0.19614g 0.1±0.01 78±1 0.13±0.01 5/18 0.19614g 0.03±0.00 133±1 0.02±0.00 8/11 0.19614g 0.1±0.00 49±1 0.2±0.01 5/11 Poaceae Aristida 0.19614g 0.03±0.00 153±5 0.02±0.00 8/5 queenslandica var. 0.19614g 0.07±0.01 61±0 0.11±0.02 queenslandica 4/9 Poaceae Aristida vagans 0.19614g 0.03±0.01 38±2 0.07±0.03 5/10 0.19614g 0.03±0.00 127±1 0.03±0.00
* Transition metal concentration between the two transport systems HATS and LATS (mg plant available metal.kg-1 soil DW) taken from the literature (see Table 3-4) and converted from µM into mg metal.kg-1 soil DW, assuming water content of the soil samples for plant uptake is 15% (w/w) (Fredlund and Xing, 1994). -1 †Plant available metal in soil (0.01 M CaCl2 extraction) (mg plant available metal.kg soil DW). References: a(Guerrier, 1978); b(Bowen, 1969); c(Bowen, 1981); d(Wiren et al., 1995); e(Pedas et al., 2005); f(Aschmann and Zasoski, 1987); g(Lasat et al., 1996).
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Our field survey of plant available concentrations of Al, Cu, Fe, Mn, Ni and Zn in soils confirmed that the proposed metal uptake conceptual framework (Figure 3-1) is valid for determination of the distribution of metals within plants. Plant available Al, Fe, Mn and Ni at all sites, Cu at two sites and Zn at one site were above the transition concentrations between high- and low-affinity transport systems. Therefore, distributions of Al, Fe, Mn and Ni within plants were expected to be facilitated by LATS. For these metals, uptake slope, intercept and coefficient of determination (R2) were determined (Table 3-6). It is relevant that linear regressions explained a high proportion of the variation in tissue metal concentration for almost all associations. In contrast, plant available concentrations of Cu and Zn at most sites were below transition concentration between transport systems, so that uptake of these metals into plants was carried out predominantly by HATS.
Therefore, accumulation of Cu and Zn in plant parts was measured by Vmax and Km, and coefficients of determination (R2) were described for HATS using the Lineweaver-Burk method (Table 3-7).
For example, the slope of shoot/leaf vs. soil concentration of a well-known Ni-hyperaccumulator (Pimelea leptospermoides) was 78.03 mg.kg-1 plant DW.mg-1 available metal.kg-1 soil DW, which was around seven-fold higher than the slope for the root (10.68 mg.kg-1 plant DW.mg-1 available metal.kg-1 soil DW) (Table 3-6). This means that P. leptospermoides tends to accumulate Ni in its upper parts (shoot/leaves) much more than in its roots. Therefore, the species is a “shoot/leaf tolerator”. This is confirmed by the species’ Ni concentration where the shoot and leaves of the species collected from three different sites contained between 6.55 and 2304 mg Ni.kg plant DW-1 compared to between 4.16 and 332.45 mg Ni.kg plant DW-1 in the corresponding roots (Table 3-8).
In contrast, the slope of Cu uptake in leaves and combined leaf and stem samples of Melinis repens vs. substrate concentration was one-quarter of the corresponding slope for roots, indicating that this species tends to accumulate Cu in its roots and is a Cu-root tolerator. This is consistent with the Cu concentrations in plant parts, where leaves and combined leaf and stem samples of M. repens contained only 2.10±0.27 and 5.84±0.83 mg Ni.kg plant DW-1, respectively compared to 9.03±0.5 and 23.01±0.63 mg Ni.kg root DW-1, in the corresponding roots (Table 3-8).
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Table 3-6 LATS metal uptake parameters derived from field collected plant species. Parameter values are given as the mean for each species across all sites. Transition Field LATS parameters† Plant metal tolerance Plant concn * Metal Family Plant Species characteristics part Intercept (mg.kg-1) Slope R2 (mg.kg-1) Al Myrtaceae Corymbia xanthope No inf L(n+o) 0.1011d 29.141 8.6059 0.7868 L (s) 0.1011d 13.593 14.241 0.7795
Eucalyptus fibrosa No inf L 0.1011d 23.478 1.0427 0.7304
Eucalyptus New Al-hyperaccumulator Sh, L 0.1011d 17.203 14.206 0.5468 tereticornis [L]=4733 mg kg-1 **a L, Poaceae Aristida vagans No inf 0.1011d 31.74 90.498 0.8603 L+St Sorghum nitidum New Al-hyperaccumulator L 0.1011d 246.55 227.22 0.7556 forma aristatum [L]=1594 mg.kg-1 **a Themeda triandra No inf Sh, L 0.1011d 9.5844 104.81 0.6024 Cu Poaceae Melinis repens No inf L 0.9532e; f 1.6794 0.4515 0.8027 R 0.9532e; f 6.3537 2.7007 0.9865 Fe Poaceae Panicum mitchellii No inf Sh, L 0.2513g 66.49 362.95 0.6756 Sorghum nitidum New Fe-accumulator L 0.2513g 101.69 212.32 0.8490 forma aristatum [L]=1043 mg.kg-1**a
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Table 3-6 contd. Transition Plant metal Field LATS parameters† concn * Metal Family Plant Species tolerance Plant part Intercept characteristics (mg.kg-1) Slope R2 (mg.kg-1) Mn Cyperaceae Scleria mackaviensis No inf L 0.0010h 19.522 7.6135 0.9293 L (s) 0.0010h 3.3637 80.043 0.8934
R 0.0010h 19.094 161.38 0.56
Mimosaceae Acacia leptostachya Mn-tolerantb Sh, L 0.0010h 3.3963 8.3752 0.6261 Acacia salicina No inf Sh, L 0.0010h 17.991 133.04 0.9910 Myrtaceae Corymbia xanthope No inf L 0.0010h 14.36 0.2813 0.8128 L (s) 0.0010h 9.1225 24.588 0.9952 Eucalyptus No inf Sh, L 0.0010h 124.39 35.516 0.9640 tereticornis
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Table 3-6 contd. Transition Plant metal Field LATS parameters† concn * Metal Family Plant Species tolerance Plant part Intercept characteristics (mg.kg-1) Slope R2 (mg.kg-1) Mn Poaceae Aristida No inf L 0.0010h 8.5334 19.411 0.9196 queenslandica var. R 0.0010h 35.185 66.616 0.9914 queenslandica Cymbopogon No inf Sh, L 0.0010h 5.8396 36.571 0.8465 refractus Heteropogon No inf Sh, L 0.0010h 3.3584 59.967 0.8788 contortus Sh, L (s) 0.0010h 27.51 68.123 0.6909
R 0.0010h 54.151 25.262 0.9910
Panicum mitchellii No inf Sh, L 0.0010h 5.18 17 23.124 0.9195
Sorghum nitidum No inf L 0.0010h 15.718 38.271 0.7156 forma aristatum R 0.0010h 130.88 14.8 0.9361
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Table 3-6 contd. Transition Field LATS parameters† Plant metal tolerance Plant concn * Metal Family Plant Species characteristics part Intercept (mg.kg-1) Slope R2 (mg.kg-1) Pimelea Mn Thymelaeaceae No inf Sh, L 0.0010h 2.2261 91.13 0.9072 leptospermoides R 0.0010h 1.6264 43.194 0.701
Cyperaceae Scleria Ni No inf R 0.088i 42.648 37.858 0.7746 mackaviensis Mimosaceae Acacia leptostachya No inf Sh, L 0.088i 1.7253 3.9785 0.6713
Myrtaceae Corymbia xanthope New Ni-accumulator L (n) 0.088i 4.4143 17.703 0.9103
[L]=115 mg.kg-1**a
Eucalyptus fibrosa No inf L(n+o) 0.088i 2.3518 1.2029 0.6808
Cymbopogon Poaceae No inf R 0.088i 6.1507 0.037 0.9997 refractus Heteropogon Ni-accumulatorc L 0.088i 0.3423 2.0938 0.7656 contortus L (s) 0.088i 2.3294 0.0816 0.9991
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Table 3-6 contd. Transition Field LATS parameters† Plant metal tolerance Plant concn * Metal Family Plant Species characteristics part Intercept (mg.kg-1) Slope R2 (mg.kg-1) Pimelea Known Ni- Ni Thymelaeaceae leptospermoides hyperaccumulator Sh, L 0.088i 78.026 0.1083 0.9942 [S]=2304 mg.kg-1**a R 0.088i 10.676 0.6773 0.998
Zn Mimosaceae Acacia salicina No inf Sh, L 0.1961j 72.532 4.9496 0.9947
* Transition concentration between transport systems (mg plant available metal.kg-1 soil DW). † Low-affinity transport system parameters: Slope, ([mg.kg-1 plant DW]/[mg available metal.kg-1 soil DW]); Intercept, (mg metal.kg-1 plant DW); R2, coefficient of determination. Field LATS combines both constitutive and inducible uptake via the high and low affinity transport systems. Definitions: ** Leaf, shoot, root [L,S,R] Al concentration >1000 mg Al.kg-1 [L=1000] foliar threshold value for hyperaccumulator; or Fe concentration >500 mg Fe.kg-1 [L=500] foliar threshold value for accumulator; or Ni concentration >100 mg Ni.kg-1 [L=100] foliar threshold value for accumulator, or >2000 mg Ni.kg-1 [L=2000] foliar threshold value for hyperaccumulator. Abbreviations: L, leaf; n, new growth; o, old growth; s, senescent; Sh, shoot; St, stem; R, root; No inf, no information available in the literature. References: a(Baker, 1981); b(Batianoff et al., 2000); c(Muhammad et al., 2013); d(Guerrier, 1978); e(Bowen, 1981); f(Bowen, 1969); g(Wiren et al., 1995); h(Pedas et al., 2005); i(Aschmann and Zasoski, 1987); j(Lasat et al., 1996).
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Table 3-7 HATS metal uptake parameters derived from field collected plant species. Parameter values are given as the mean for each species across all sites. Plant metal Transition Metal Family Plant Species Plant part Field HATS parameters† tolerance concn* characteristics Vmax Km (mg.kg-1) R2 (mg.kg-1) (mg.kg-1) Cu Myrtaceae Eucalyptus fibrosa No inf R 0.9531a; b 6.5963 0.2058 0.7734 Poaceae Aristida vagans No inf L, L+St 0.9531a; b 13.0718 0.0967 0.6405
Themeda triandra No inf Sh, L 0.9531a; b 7.7881 0.0404 0.7813 Zn Myrtaceae Eucalyptus fibrosa No inf L 0.1961c 12.8205 0.0730 0.7003 Poaceae Aristida queenslandica var. No inf R 0.1961c 312.50 0.4062 0.8724
queenslandica Aristida vagans No inf L, L+St 0.1961c 27.5482 0.0110 0.9869
* Transition concentration between transport systems (mg plant available metal.kg-1 soil DW). -1 -1 2 † High-affinity transport system parameters: Vmax, (mg metal.kg plant DW); Km, (mg metal.kg soil DW); R , coefficient of determination using the Lineweaver-Burk method. Field HATS combines both constitutive and inducible uptake. Abbreviations: L, leaf; Sh, shoot; St, stem; R, root; No inf, no information available in the literature. References: a(Bowen, 1981); b(Bowen, 1969); c(Lasat et al., 1996).
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Another grass species, Heteropogon contortus, accumulated Mn mainly in its roots, with the highest uptake slope (54.15 mg.kg-1 plant DW.mg-1 available metal.kg-1 soil DW), followed by senescent shoot and senescent leaves (27.51 mg.kg-1 plant DW.mg-1 available metal.kg-1 soil DW), while the lowest slope value was observed in shoots and leaves (3.36 mg.kg-1 plant DW.mg-1 available metal.kg-1 soil DW). With the exception of material collected from Site 4/5, Mn concentration in roots of H. contortus (493.5±24.5 mg Mn.kg root DW-1) was between 1.60 and 5.55 times higher than the Mn concentration in senescent leaves and leaves, respectively (Table 3-8). Similarly, Aristida queenslandica var. queenslandica was also observed to store Mn primarily in its roots rather than its aerial part. Its root uptake slope was 4 times higher than that for its leaves, and Mn concentrations in roots of this species collected from two different sites were also 4-fold higher than those in the leaves (Table 3-8).
Our field data also showed that the low-affinity uptake intercept parameter is generally less sensitive than the slope parameter to describe plant metal uptake. For example, the Ni- hyperaccumulator P. leptospermoides showed similar uptake intercepts (less than 1 mg Ni.kg-1 plant DW) for both aerial and belowground parts although the species accumulated high concentrations of Ni in its shoot (Table 3-8).
At low external Cu and Zn concentrations (below their corresponding transition concentration between transport systems), all the species studied had high Vmax but low Km (Table 3-7). Aerial parts of Themeda triandra and Aristida vagans grown on low external Cu concentrations had Vmax -1 -1 -1 and Km of 7.79 mg Cu.kg plant DW and 0.04 mg Cu.kg soil DW and 13.07 mg Cu.kg plant DW -1 and 0.10 mg Cu.kg soil DW , respectively. Similarly, roots of Eucalyptus fibrosa had a Vmax of -1 -1 6.60 mg Cu.kg plant DW and a Km of 0.21 mg Cu.kg soil DW . For Zn, leaves and combined leaves and stems of A. vagans and leaves of E. fibrosa had high Vmax (27.55 and 12.82 mg Zn.kg -1 -1 plant DW ) and low Km (0.01 and 0.07 mg Zn.kg soil DW ), respectively. Roots of A. -1 queenslandica var. queenslandica had very high Vmax (312.50 mg Zn.kg plant DW ) and a Km of 0.41 mg Zn.kg soil DW-1.
With limited field data from sites with low available metal concentrations (often either only whole shoots or leaves or roots), we were not able to classify the species’ accumulation characteristics (i.e. tolerator type) based on our model parameters for high-affinity metal transport systems. Nonetheless, based on a study conducted by Kozhevnikova et al. (2014) on accumulation and distribution of Zn in ruderal plants Lepidium ruderale and Capsella bursa-pastoris growing in solution with Zn available concentration from 2 to 20 μM (below the transition Zn concentration
128 between transport systems), we calculated their Vmax and Km. Both species had very high root Vmax, 100,000 and 10,000 mg Zn.kg plant DW-1 for L. ruderale and C. bursa-pastoris, respectively, -1 compared to shoot Vmax which were 1429 and 385 mg Zn.kg plant DW , respectively. Thus, both species tended to accumulate Zn more rapidly in their roots than in shoots, a phenomenon which was confirmed by Zn concentrations in roots and shoots of both species. These species are classified as Zn-root tolerators according to our conceptual framework, confirming its validity.
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Table 3-8 Total metal concentration of each metal analysed via ICP in plant parts growing in mine sites and serpentine areas. Values are given as the mean ± standard errors (n=3). Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Al Myrtaceae Corymbia xanthope 1/7 72.69±2.89 42.92±6.12 2/1 11.54±3.90 15.47±0.51 5/15 71.08±17.34 Myrtaceae Eucalyptus fibrosa 1/4 10.18±5.26 1/5a 66.65±13.32 5/18 22.96±0.12 8/11 88.70±15.12 Myrtaceae Eucalyptus tereticornis Mor-1 4733±2648 4/3 39.04±2.87 7/2 31.36±7.75 Poaceae Aristida vagans 4/9 366.5±37.4 5/10 66.29±23.58s Poaceae Sorghum nitidum 4/7 354.7±25.9 6/3 1594±178 Poaceae Themeda triandra 1/11 120.4±9 8/6 152.3±5.4
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1)
Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root
Cu Myrtaceae Eucalyptus fibrosa 1/4 2.96±0.39 5/18 1.76±0.18 Poaceae Aristida vagans 4/9 11.19±1.22 5/10 8.22±0.52s Poaceae Melinis repens WTP239 2.1±0.27 9.03±0.50 WTP234 5.84±0.83s 23.01±0.63 Poaceae Themeda triandra 1/11 4.65±1.11 8/6 6.48±0.32 Fe Mimosaceae Acacia salicina WTP237 88.79±7.76 Mor-5 128.9±17.3 Myrtaceae Eucalyptus tereticornis Mor-1 190.5±25.6 4/3 46.09±3.76 7/2 49.19±22.83 Poaceae Panicum mitchellii 2/10 882.5±88.1 8/12 323±18.5 Poaceae Sorghum nitidum 4/7 356.9±27.9 6/3 1043±125
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Mn Cyperaceae Scleria mackaviensis 2/7 10.54±0.05 164.0±8.7 4/8 41.19±0.63 86.35±4.96 196.6±8.2 5/12 61.57±2.31 98.5±8.0 8/8 125.9±6.7 Mimosaceae Acacia leptostachya 1/2 57.95±1.82 1/3a 56.14±5.18 1/3b 47.68±3.87 2/4 19.88±0.95 5/9 29.72±1.31 Mimosaceae Acacia salicina WTP237 1009±24 Mor-5 172.7±19.5 Myrtaceae Corymbia xanthope 1/7 220.9±9.3 153.4±4.07 2/1 45.15±2.72 34.89±2.48 5/15 34.5±3.51 Myrtaceae Eucalyptus tereticornis Mor-1 1593±7.72 4/3 244±7.5 7/2 129.6±2.6
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Mn Poaceae Aristida queenslandica 5/11 42.88±0.5 167.8±0.6 8/5 157.2±11.4 630.5±9.6 Poaceae Cymbopogon refractus 1/10 87.87±5.83 8/4 60.39±1.90 Poaceae Heteropogon contortus 1/12 88.94±1.31 309.2±66.4 493.5±24.5 4/5 63.22±4.9 90.96±1.37s 76.6±3.5 Poaceae Panicum mitchellii 2/10 25.18±0.86 8/12 61.69±4.96 Poaceae Sorghum nitidum 4/7 67.54±1.84 268.6±9.7 6/3 47.41±3.62 98.59±9.08 Thymelaeaceae Pimelea leptospermoides 3/1 129.6±2.2 63.62±6.46 5/6 97.79±2.42 45.8±10.65 8/1 97.91±0.14 54.14±11.7
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Ni Cyperaceae Scleria mackaviensis 2/7 140.9±7.7 4/8 25.03±0.61 5/12 305.2±17.9 8/8 1.72±0.22 Mimosaceae Acacia leptostachya 1/2 71.15±0.62 1/3a 22.31±2.03 1/3b 24.73±0.62 2/4 3.96±0.12 5/9 28.86±2.69 8/9 7.15±1.71
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Ni Myrtaceae Corymbia xanthope 1/6 114.6±2.6n 2/1 14.61±0.17 5/15 56.9±2.67 Myrtaceae Eucalyptus fibrosa 1/4 11.68±0.32 1/5a 38.31±2.98 1/5b 65.26±3.69o 2/2 42.52±0.06 5/18 0.52±0.06 5/18 1.32±0.23n 8/11 0.66±0.05 Poaceae Cymbopogon refractus 1/10 139.3±5.6 8/4 1.64±0.6 Poaceae Heteropogon contortus 1/12 4.16±0.46 20.27±6.09 4/5 49.28±2.30 8/7 2.25±0.2 1.24±0.19
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Table 3-8 contd. Metal concentration in plant part (mg metal.kg plant DW-1) Metal Family Plant Species Site ShootS Leaves Senescent leavesS Root Ni Thymelaeaceae Pimelea leptospermoides 3/1 2304±71.4 332.5±19.3 5/6 349.0±12.9 45.6±4.92 8/1 6.55±0.39 4.16±0.44 Zn Mimosaceae Acacia salicina WTP237 631.6±28.5 Mor-5 17.94±2.04 Myrtaceae Eucalyptus fibrosa 1/4 6.86±0.27 1/5a 7.46±0.65 5/18 3.56±0.09 8/11 6.59±0.87 Poaceae Aristida queenslandica 5/11 18.29±0.34 8/5 41.62±0.88 Poaceae Aristida vagans 4/9 14.19±0.81 5/10 20.89±0.44s Abbreviations: Aristida queenslandica: Aristida queenslandica var. queenslandica Sorghum nitidum: Sorghum nitidum forma aristatum
Symbols : n : new leaves o : old leaves S : or stem + leaves
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3.5 Discussion The study proposes a new conceptual framework for characterization of plant responses to metals in soil based on metal transporter kinetic parameters. This framework is the first to use these parameters to describe the distribution of metals in plant parts at both low and high external plant available metal concentrations.
The proposed plant uptake framework advocates that plant available metal concentration in soil is the most relevant form to use when studying soil-plant metal relationships. Based on our study of plant and soil samples collected from metal-enriched areas, only a small fraction of total metal in soil was available for plant uptake. Concentrations of Mn, Ni, Fe, Al and Cu (some samples) in plant parts of the species collected correlated better with plant available metals in soils rather than total metal concentrations in soils. This finding conforms with previous studies which have shown that plants absorb heavy metals that are present as soluble components in the soil solution and thus are readily available for plant roots (Chibuike and Obiora, 2014, Garcia-Gomez et al., 2014, Saha et al., 2017). In addition, bioavailability of metal rather than total metal concentration governs the physiological and toxic effects of a metal on biological systems (Jamali et al., 2009, Olaniran et al., 2013, Sahito et al., 2015, Saha et al., 2017). Similarly, Lu et al. (2017b) showed that there was no significant correlation between total Co, Cr, Cu, Mn, Ni, Pb and Zn concentrations in soils and corn grain. Fernández-Calviño et al. (2017) found that the environmental risks of Cu pollution are more closely related to its available form than to the total soil Cu concentration. Similarly, Roca et al. (2017) showed that the extractable fraction of metals in soils can be related to the bioaccumulation factor for the identification of native plants with the capacity to take up high concentrations of metals from soils and accumulate them in their tissues. Other studies have also shown that soil available metal concentration is the most appropriate basis for examining plant metal uptake (Jamali et al., 2006, Zhang et al., 2006, Massas et al., 2013, Wójcik et al., 2014, Mukhopadhyay et al., 2017). Thus, we conclude that plant available form of metal in soil is a more relevant descriptor than its total form to use in plant-soil metal relationship models.
The Baker (1981) framework has been used widely for nearly 40 years to describe plant response to metals in soils (Gabrielli et al., 1991, Brunner et al., 2008, Fernando et al., 2009, Feng et al., 2011, Callahan et al., 2012, Iori et al., 2013, van der Ent, 2013, Fuentes et al., 2014, Nkrumah et al., 2016) and to determine whether metal tolerant plants belong to one of the four proposed categories based on their accumulation potential in their aboveground tissues, namely metal accumulators, hyperaccumulators, indicators and excluders. According to this framework, accumulators and hyperaccumulators would display a hyperbolic response, indicators a linear response, while in excluders, plant uptake remains constant and low over a wide range of soil concentrations (Baker, 137
1981). Our proposed conceptual framework predicts that plants growing on areas where external plant available concentrations of metals are higher than their respective transition concentration between transport systems display linear metal uptake relationships. Our field study focusing on metal-enriched areas (serpentine and mine sites) showed that this proposition was confirmed for Mn, Ni, Fe, Al and Cu (some samples) in all soil samples tested. We found linear relationships between plant available metal concentrations in soils and the metal concentrations in plant parts; e.g., P. leptospermoides collected from 3 different serpentine areas (Sites 3/1, 5/6 and 8/1) accumulated greater amounts of Ni than Mn in the shoot/leaves, as shown by the LATS slope for these metals of 78.03 and 2.23, respectively (Table 3-6). The finding of this linear response differs from Baker’s original hypothesis, which proposed non-linear (hyperbolic) relationships between heavy metal concentrations in tolerant plants grown on soils from low to high total soil metal concentrations rather than plant available metal concentrations, except for metal indicators (Baker, 1981). Recent work (e.g., Nkrumah et al. (2016)) refers to extractable metal concentrations in soils, but retains the original concept of a hyperbolic hyperaccumulator response. While for the above- mentioned metals (found in soils at plant available concentrations higher than the transition concentration between transport systems), the linear plant metal uptake was facilitated by LATS, other metals such as Cu and Zn were found to be present in most of the soils tested at plant available concentrations lower than their transition concentration between transport systems, and in these cases, the plant response was hyperbolic and presumably facilitated by HATS.
Our study identified four sensitive metal transporter kinetic parameters which can be used to determine the distribution of metals in any plant parts from low or high soil metal concentrations and to classify plant species into shoot/leaf, root, or shoot/leaf-root tolerators. Maximum uptake
(Vmax) and Km were used to measure metal uptake by plants via HATS under saturable, low metal concentration conditions, while the linear slope (the intercept was found not to be a sensitive parameter) describes uptake at higher concentrations mainly via LATS. The framework also recognises that the transition concentration between transport systems is an important parameter as it determined whether metal uptake occurs at saturable or linear rates.
The HATS parameters (Vmax and Km) have been used largely to measure the influx and uptake efficiency from solutions of some metal/loids into plants, such as low soluble concentrations of Mn (Hassan, 1977, Pedas et al., 2005), Zn (Lasat et al., 1996, Shankhdhar et al., 1999, Hacisalihoglu et al., 2001, Rosolem et al., 2005, Nie et al., 2017) and Fe (Wiren et al., 1995), and also to study responses of tolerant/non-tolerant or hyperaccumulator/non-accumulator plants with respect to low soluble concentrations of Zn (Lasat et al., 1996); Cu (Landi and Fagioli, 1983, Strange and Macnair, 1991) and As (Meharg and Macnair, 1992a, Meharg and Macnair, 1992b, Abedin et al., 2002, Zhao 138 et al., 2002a, Poynton et al., 2004). Neither Vmax nor Km have been utilized as plant metal influx parameters to describe plant metal uptake from soils, although they have been used to measure ammonium and nitrate uptake in the laboratory (Goyal and Huffaker, 1986, Wijk and Prins, 1993, Kronzucker et al., 1998) and to successfully predict nitrate uptake in field conditions (Faure- Rabasse et al., 2002, Malagoli et al., 2004).
In contrast to the HATS parameters, the LATS parameters, slope and intercept, have received less attention. Although there are many studies on plants from metal-enriched areas or exposed to high external metal concentrations, especially Ni (Reeves et al., 1996, Robinson et al., 1997, Robinson et al., 2003, Jaffré et al., 2013, van der Ent et al., 2015), Mn (Bidwell et al., 2002, Xue et al., 2004, Fernando et al., 2009, Liu et al., 2010, Yang et al., 2013c, Liu et al., 2014) and As (Codling and Ritchie, 2005, Flores-Tavizón et al., 2003, Chang et al., 2009, Chandra et al., 2018), LATS slope and intercept have not been utilized for determination of metal uptake by plants. Limited studies have been conducted on these parameters, specifically slope, but mainly on element influxes into plants from solutions (Kronzucker et al., 1995, Lasat et al., 1996, Cerezo et al., 1997, Cohen et al., 1998, Li et al., 2011b).
3.6 Conclusion We suggest a new terminology for metallophytes (i.e. tolerators) that matches our proposed conceptual framework which describes plant responses to metal availability in soils based on metal transporter kinetic parameters. The framework was validated using independent datasets from both field surveys and the literature, and is valid for determination of the distribution of metals in all plant parts, and plants growing in low to high plant available metal concentrations in soils. This conceptual framework may be a useful tool for selecting suitable metal tolerators for specific phytoremediation purposes, and may be applied to non-metal elements or ions.
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Chapter 4. Assessing germination characteristics of Australian native plant species in metal/loid solution
The chapter has been accepted and published as a research article in a high ranking refereed journal, Journal of Hazardous Materials, 2019 with the following detail:
Guterres, J., Rossato, L., Doley, D., Pudmenzky, A., Bee, C., Cobena, V. 2019. Assessing germination characteristics of Australian native plant species in metal/metalloid solution. Journal of Hazardous Materials, Vol. 364C, p.173-181.
For uniformity of the thesis, some changes in the formatting have been made including omitting the ‘article highlights’ and inclusion of the references in the complete thesis reference list. All the experiments described in the chapter/paper were performed by candidature with the assistance of Mr Victor Cobena (a 2007 UQ Master student) and Ms Christina Bee (a 2009 UQ Summer Scholarship student) and under the supervision of the advisors Dr Laurence Rossato and Dr David Doley, and Dr Alex Pudmenzky. Data analysis and preparation of figures described in the paper were performed by the candidature under the supervision of the advisors Dr Laurence Rossato and Dr David Doley. The Matlab data analysis was performed by Dr Alex Pudmenzky and experiment photographs were taken by Drs Alex Pudmenzky and Laurence Rossato.
4.1 Abstract This study investigated tolerance of the Australian native grass species Astrebla lappacea, Themeda australis, and Austrostipa scabra and a tree species Acacia harpophylla, to different concentrations of arsenic As(V) (13.34-667.36 μM), Cu2+ (0.5-200 μM), Zn2+ (9-500 μM), Mn2+ (8-10240 μM) and Pb2+ (240-9600 μM) in single solutions in germination experiments. Metal/loid tolerance indicators used were maximum germination percentage (Gmax), mean germination time (MGT), and radicle and shoot tolerance indexes (RTI & STI). Radicle tolerance index was the most sensitive indicator of metal tolerance in germinating seeds. All native species were highly tolerant to the metal/loids tested, however, they showed different metal toxicity thresholds and levels of tolerance based on RTI as a metal tolerance indicator during germination. Overall, all four species could be classified as metallophytes, confirming their current suitability for and established use in mine site rehabilitation. This work may also serve as a basis for future studies on metal/loid tolerance of other plant species during germination.
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Keywords: germination, Australian native plant species, metal/loid phytotoxicity threshold, tolerance indicator, radicle tolerance index
4.2 Introduction Metal accumulation in the environment has become a widespread problem as toxic soluble metals have been released into the biosphere through both natural processes and human activities, such as mining and agriculture (Kabata-Pendias, 2001, Sharma and Dietz, 2008, Nagajyoti et al., 2010). Although low concentrations of metals such as manganese (Mn), copper (Cu), zinc (Zn), molybdenum (Mo) and nickel (Ni), are essential to microorganisms, plants, and animals (Chehregani and Malayeri, 2007, Rascio and Navari-Izzo, 2011), at high concentrations, they have strong toxic effects and are environmental threats (Munzuroglu and Geckil, 2002, Wuana et al., 2010, Misaelides, 2011).
A common objective of dealing with waste metals in mining environments is to revegetate waste storage facilities that may contain toxic concentrations of these metals (Lottermoser, 2010b). Where metal properties are highly problematic, they may be covered with impermeable barriers and layers of benign material deep and stable enough to be secure almost indefinitely (Zhang et al., 2017b). While the design of these facilities may aim to intercept and evaporate all of the incident rainfall and prevent the release of soluble forms of these metals (Bell, 2001, Rock, et al., 2012), this is rarely achieved in Australian coal and metalliferous mines (Glenn et al., 2014, Lamb et al., 2015). As a result, various remediation procedures are applied to the surfaces of tailings dams and waste heaps in order to establish vegetation (Mendez and Maier, 2008a, Tongway and Ludwig 2011, Erickson et al., 2016). However, where sufficient cover materials of suitable quality are not available, plant establishment may be compromised by toxic concentrations of metals in the surface material unless economical soil amendment can be achieved. The application of polymer particles to surface soils is one such amendment that has enabled the germination and establishment of plant species on otherwise toxic substrates (Rossato et al., 2011, Guterres et al., 2013). However, it is difficult to totally eliminate soluble metals from these environments, and it is important to understand the tolerance to elevated metal concentrations of plant species that may be selected for revegetation work.
Trees and grasses are commonly used for revegetation of degraded and/or metal contaminated land (Mendez and Maier, 2008a, Mahar et al., 2016, Rahman et al., 2016). In Australia, many native plant species have been used in mine site rehabilitation as they are tolerant to drought (Huxtable, 2003, Erickson et al., 2016, Nirola et al., 2016), low soil fertility (Mendez and Maier, 2008a, Lamb et al., 2010), salinity (Mendez and Maier, 2008a) and poor soil structure (Lamb et al., 2010, Lamb 141 et al., 2015, Macdonald et al., 2017). Native plants are also preferred over exotic (i.e. non-native) species (Lamb et al., 2010, Guterres et al., 2013, Erickson et al., 2016) as they do not present the risk of becoming invasive and reducing local plant biodiversity (Keeling and Werren, 2005, Barrutia et al., 2011).
Metal tolerance has been reported widely for plants native to temperate regions on the basis of field observations (Smith and Bradshaw, 1979, Shaw, 1989) and for crop and pasture species on the basis of experimental solution culture studies (Wong and Bradshaw, 1982, Kopittke et al., 2010). There have been comprehensive listings for mine sites in different areas of Australia of native species generally (Erickson et al., 2016, Grant et al., 2002, Archer and Caldwell, 2014) and grasses in particular (Huxtable, 2003) that can be recommended for revegetation work. These valuable studies of the outcomes of plant adaptation did not always describe quantitative responses of plant growth to plant-available metal concentrations in the soils.
Experimental studies of the effects of plant available metal concentrations on the growth of seedlings in soil or culture solution have been conducted by Reichman (2001) and Reichman et al. (2004, 2006). Germination is often the stage of plant development that is most sensitive to environmental stress (Souza and Cardoso, 2000) so it is important to understand its response to plant-available metal concentrations in the soil solution. This aspect of mined land restoration has been considered for some situations in southern Australia (Nirola et al., 2016, Lamb et al., 2010), but not in detail for tropical environments or for mining situations in which site amelioration by techniques such as polymer particle application may be advantageous (Guterres et al., 2013).
The present study was undertaken to establish the limits within which selected semi-arid Australian native plant species including grasses (Astrebla lappacea, Austrostipa scabra and Themeda australis) and a tree (Acacia harpophylla) might be able to germinate on different plant-available concentrations of As(V), Cu2+, Zn2+, Mn2+ and Pb2+ commonly found in mine wastes.
4.3 Materials and methods 4.3.1 Plant materials Grass and tree species native to Australia were chosen for the study. Seeds of Astrebla lappacea var. yanda and Themeda australis were obtained from AustraHort Pty Ltd (Queensland, Australia) and from Native Seeds Pty Ltd (Cheltenham, Victoria, Australia), respectively. Seeds of A. lappacea were hand-collected between March and April 2006 from Central Queensland (Australia), and in March 2007 from Walgett (NSW, Australia), and stored for the 4-6 months required for after-ripening (Ralph, 2003). Seeds of T. australis were collected between March and April 2006
142 from Penrith (NSW, Australia), and in December 2007 from Coonabarabran (NSW, Australia). T. australis seeds collected from this area of south-east Australia will generally only have a small percentage of seeds that are dormant (Ralph, 2003). Seeds of Acacia harpophylla were obtained from AustraHort Pty Ltd (Queensland, Australia), and seeds of Austrostipa scabra from Native Seeds Pty Ltd (Cheltenham, Victoria, Australia). Seeds of A. harpophylla were collected from Roma (Queensland, Australia) in January 2006 and did not require any pre-treatment to germinate (Arnold et al., 2014). Seeds of A. scabra were collected in January 2008 from Ardlethan (NSW, Australia), easily accommodating the storage period of four months recommended by Ralph (2003) to overcome dormancy. Given that A. lappacea and A. scabra seeds were stored for longer than the recommended time required to overcome dormancy, and T. australis and A. harpophylla seeds were not expected to display dormancy, no pre-treatment was applied to seeds. All seeds collected were stored air-dry at 15˚C before the start of germination experiments.
4.3.2 Chemicals . Water soluble metal and metalloid compounds were used throughout. As(V) (Na2HAsO4 7H2O) was 2+ . 2+ purchased from Sigma-Aldrich (Germany), Cu (CuCl2H2O) and Zn (ZnCl2) from Chem-Supply 2+ . Pty Ltd (South Australia, Australia), Mn (MnCl2 4H2O) from Asia Pacific Specialty Chemicals 2+ Ltd (NSW, Australia) and Pb (PbCl2) from Lomb Scientific (Aust.) Pty Ltd (Coopers Plains, Queensland, Australia). pH of the metal/loid solutions was not measured, but we did not observe any signs of metal/loid precipitates in the solution at any stage of the experiment. Metal/loid solutions were also tested via ICP before use to confirm metal/loid concentration.
4.3.3 Germination experiments Effects of various concentrations of As(V), Cu2+, Zn2+, Mn2+ and Pb2+ on the germination of the selected native species were evaluated using single metal solutions freshly prepared in sterilized deionized water (SDIW). Two sets of experiments were conducted and metal/loid concentrations tested are presented in Table 4-1.
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Table 4-1 Metal/loid concentrations tested on seed germination of native plant species Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla. Metal/loid concentration (μM) Plant species As(V) Cu Zn Mn Pb A. lappacea 13.34 0.5 9 8 240 T. australis 66.73 10 25 128 480 A. scabra 133.47 40 100 2048 960 266.94 80 200 4096 1920 400.41 200 500 10240 3840 533.88 - - - 4800 667.36 - - - 9600 A. harpophylla 1st experiment 13.34 0.5 9 8 - 66.73 10 25 128 - 133.47 40 100 2048 - 2nd experiment 133.47 - - 2048 - 266.94 - - 4096 - 667.36 - - 10240 -
Prior to germination experiments, seeds were surface-sterilised in 20% sodium hypochlorite for 10 min and then rinsed three times in SDIW for one min. All seed manipulations and examinations were then conducted in a laminar flow cabinet (AES Environmental, CLYDE-APAC & Contamination Control Laboratories, Australia) to minimise fungal and bacterial contamination throughout the experiment. Disposable Petri dishes (9 cm) each containing two 84 mm filter papers discs (Advantec) autoclaved for 15 min at 121˚C prior to experiments (Tomy ES-315 Autoclave, Tokyo, Japan) were used for germination. For all species, 25 seeds were placed in each Petri dish. Five (A. lappacea and A. harpophylla), eight (A. scabra), or either five, eight or ten (T. australis) replicate dishes were used for each species depending on seed viability. Ten mL of metal solution was added to each treatment Petri dish and 10 mL of SDIW was used for the control.
The Petri dishes were closed, sealed with Parafilm and placed in a transparent plastic zip resealable bag to reduce water evaporation. They were then incubated in a controlled temperature cabinet at optimum conditions of day and night temperatures of 30.2±0.1˚C and 25.2±0.1˚C (Sindel et al., 1993, Grice et al., 1995, Reichman et al., 2006) with 12:12 hours white light and dark conditions, and with moisture non-limiting. The bags and dishes were opened daily for measurements and to maintain oxygenation.
Newly germinated seeds were counted and removed daily for grasses, and either daily (experiment 1) or at 3.08, 6.25, 9.25, 12.25, 15.41, 18.91, 21.91, 24.91, 48.91, 72.91, 103.41, 120.41 and 151.91 144 hours (experiment 2) for A. harpophylla, until the maximum germination percentage was reached i.e. after 14 days for A. lappacea and A. scabra, either after 14 or 16 days for T. australis, and after 7 days i.e. 168 hours (experiment 1) or 151.91 hours (experiment 2) for A. harpophylla.
Seeds were considered germinated when 0.5 mm (A. lappacea, A. scabra and T. australis) or 1 mm of radicle (A. harpophylla) had emerged, or alternatively, if no radicle had emerged, when 2 mm of shoot had emerged. In cases where a single diaspore contained more than one germinated caryopsis, the length of only the longest radicle or shoot was recorded and a single germinated diaspore counted.
4.3.4 Determination of metal tolerance indicators Four indicators were used to assess the tolerance of the plant species to the selected metals during germination, namely maximum germination percentage (Gmax), mean germination time (MGT), radicle tolerance index (RTI) and shoot tolerance index (STI). For grasses, indicators used to determine species’ tolerance to metal/loids were Gmax, MGT, RTI and STI (As(V) and Pb treatments only). For A. harpophylla, in experiment 1, only Gmax and MGT were calculated and experiments were monitored daily, while in experiment 2, monitoring was done hourly and included
Gmax, MGT and RTI.
4.3.4.1 Maximum germination percentage (Gmax)
Gmax was calculated as the mean of the maximum cumulative germination percentage observed in each replicate Petri dish for each treatment. In some experiments, the plumule emerged from some seeds without the prior or contemporaneous appearance of the radicle (germination was described as “with shoot only” or Gmax S), while in others, radicle emergence was the first response
(germination was described as “with radicle” or Gmax R).
4.3.4.2 Mean Germination Time (MGT) MGT was calculated using the following formula of Ellis and Roberts (1981): Dn MGT(hours/ days) (1) n Where: D = number of days (grasses) or hours (tree) since the beginning of the experiment, n = number of germinated diaspores/seeds on day/hour D
As for Gmax, mean germination time of seeds germinated “with shoot only” was defined as MGT S while germinated seeds “with radicle” was defined as MGT R.
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4.3.4.3 Radicle (RTI) and Shoot (STI) Tolerance Indices Radicle length of germinated seeds was measured from the radicle-shoot junction to the tip of the longest radicles after 3 days for the grass species and after 9 hours for the tree species A. harpophylla and expressed as RTI using the following formula (Rossato et al., 2011):
Length of the longest radicle in metal treatment RTI (%) 100 (2) Length of the longest radicle in SDIW control
Shoot length was measured from the radicle-shoot junction to the tip of the longest shoot after 3 days and expressed as shoot tolerance index (STI):
Length of the longest shoot in metal treatment STI (%) 100 (3) Length of the longest shoot in SDIW control
4.3.5 Statistical analysis Germination data were presented as the mean and standard error of five (A. lappacea and A. harpophylla), eight (A. scabra), or either five, eight or ten (T. australis) replicates. Data were tested for significant differences between treatments using one-way analysis of variance (ANOVA) at P<0.05, and mean separation performed using the Tukey’s multiple comparison test and 95% confidence interval (P<0.05) (Matlab® Statistics Toolbox, version 2007b, The Mathworks, USA) (Rossato et al., 2011).
4.4 Results
4.4.1 Effect of metal/loids on metal tolerance indicators during germination
Species’ tolerance thresholds based on maximum germination percentage (Gmax), mean germination time (MGT), radicle and shoot tolerance indexes (RTI & STI) as tolerance indicators during 2+ 2+ 2+ 2+ germination on As(V), Cu , Zn , Mn and Pb are presented in Table 4-2. Gmax, MGT and STI were not reliable indicators of metal/loids toxicity to germination in A. lappacea, T. australis, A. scabra and A. harpophylla (detailed data in Apendices II-VIII). RTI was the most sensitive indicator (section 4.4.2) of species’ metal/loid tolerance during germination.
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Table 4-2 Metal/loid tolerance levels of Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla during germination on As(V)(13.34-667.36 μM), Cu2+(0.5- 200 μM), Zn2+(9-500 μM), Mn2+(8-10240 μM) and Pb2+(240-9600 μM), based on maximum germination percentage (Gmax), mean germination time (MGT), radicle and shoot tolerance indexes (RTI & STI) as tolerance indicators.
Metal/ Germination Species' metal/loid toxicity threshold (μM) loid indicators A. lappacea T. australis A. scabra A. harpophylla
As Gmax R <13.34 >667.36 133.47 - 266.94 >667.36 MGT R >667.36 >667.36 >667.36 >667.36 RTI 400.41 - 533.88 66.73 - 133.47 >667.36 >667.36 STI >667.36 13.34 - 66.73 >667.36 nt
Cu Gmax R >200 >200 nt >40 MGT R >200 >200 nt >40 RTI >200 >200 nt nt STI nt nt nt nt
Zn Gmax R >500 >500 nt >100 MGT R >500 >500 nt >100 RTI >500 >500 nt nt STI nt nt nt nt
Mn Gmax R >10240 >10240 nt >10240 MGT R >10240 >10240 nt >10240 RTI >10240 >10240 nt >10240 STI nt nt nt nt
Pb Gmax R 1920 - 3840 >9600 4800 - 9600 nt MGT R 4800 - 9600 Inconclusive 3840 - 4800 nt RTI 3840 - 4800 <240 >9600 nt STI 4800 - 9600 240 - 480 >9600 nt 'Gmax R' stands for 'maximum germination percentage of germinated seeds with radicle'; 'MGT R' stands for 'mean germination time of germinated seeds with radicle'; 'nt' stands for 'not tested’.
4.4.2 Effect of metal/loids on radicle tolerance index (RTI)
4.4.2.1 As(V) There was no significant difference between RTI of A. lappacea in As(V) concentration compared to the control until the concentration reached 533.88 µM, where the RTI began to decrease significantly (P<0.05) and reached 9.4% of the control value at 667.36 µM As(V) (Table 4-3 and Figure 4-1). These results show that the RTI toxicity threshold was between 400.41 and 533.88 µM As(V). Consistent toxic effects of higher concentrations of As(V) suggests that RTI was responsive to the range of concentrations tested.
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Table 4-3 Radicle tolerance index (RTI) of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days or Acacia harpophylla after 9 hours on various concentrations of As(V), Cu, Zn, Mn and Pb solutions. RTI (%)* Metal/loid A. lappacea T. australis A. scabra A. harpophylla 2nd exp.
Control (SDIW) 100±0a 100±0a 100±0a 100±0a As(V) (13.34 μM) 123.1±48.3a 72±37.2a 623.2±264.2b nt As(V) (66.73 μM) 104.5±23.1a 51.0±38.7a 261.9±139.5a nt As(V) (133.47 μM) 119.6±29.0a 30.4±22.2b 94.6±37.5a 86.7±6a As(V) (266.94 μM) 61.5±10.1a 21.5±17.5b 65.5±16.3a 66.7±6.2a As(V) (400.41 μM) 39.6±18.2a 6.5±4.2b 76.8±26.4a nt As(V) (533.88 μM) 22.6±14.8b 11.1±8.3b 62.5±28.1a nt As(V) (667.36 μM) 9.4±9.4b 1.4±1.4b 63.7±27.7a 66.7±13.6a Control (SDIW) 100±0a 100±0a nt nt Cu (0.5 μM) 106.7±6.7a 66.7±33.3a nt nt Cu (10 μM) 65.3±16.9a 104.4±54a nt nt Cu (40 μM) 96.7±13.3a 61.1±2a nt nt Control (SDIW) 100±0a 100±0a nt nt Cu (40 μM) 130±8.6a 36.5±16.2a nt nt Cu (80 μM) 111±17.3a 146.1±64.1a nt nt Cu (200 μM) 64±16.1a 65.8±29.9a nt nt Control (SDIW) 100±0a 100±0a nt nt Zn (9 μM) 73.3±11.3a 58.3±14.4a nt nt Zn (25 μM) 78.7±21.7a 116.7±35.4a nt nt Zn (100 μM) 96.7±15.5a 46.1±28.2a nt nt Control (SDIW) 100±0a 100±0a nt nt Zn (100 μM) 98±14.5a 87.2±33.5a nt nt Zn (200 μM) 122.7±8.2a 64.6±25.6a nt nt Zn (500 μM) 100±13.9a 54.18±14.8a nt nt
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Table 4-3 contd. RTI (%)* Metal/loid A. lappacea T. australis A. scabra A. harpophylla Control (SDIW) 100±0a 100±0a nt nt Mn (8 μM) 101.7±9.3a 44.4±5.6a nt nt Mn (128 μM) 91.7±13.9a 64.6±45.8a nt nt Mn (2048 μM) 93.8±6.3a 109.2±34a nt nt 2nd exp. Control (SDIW) 100±0a 100±0a nt 100±0a Mn (2048 μM) 101.7±8.4a 157.7±93a nt 66.7±6.2a Mn (4096 μM) 117.7±10.3a 189.6±104.8a nt 66.7±6.2a Mn (10240 μM) 101±7.1a 89.6.4±28.4a nt 86.7±13.6a Control (SDIW) 100±0a 100±0a 100±0a nt Pb (240 μM) 137.6±27.4a 35.9±22.6b 554.3±317.4b nt Pb (480 μM) 139.6±46.8a 10.9±6.8b 258.4±163.6a nt Pb (960 μM) 152.4±38.9a 0±0b 144.0±70.4a nt Pb (1920 μM) 92.8±24.2a 49.5±34b 32.9±10.9a nt Pb (3840 μM) 48.8±25.5a 0±0b 64±35.3a nt Pb (4800 μM) 14.4±9.0b 6.3±6.3b 36.5±24a nt Pb (9600 μM) 0±0b 15.6±11.8b 35.2±14.2a nt
*Results are given as the mean and standard error of five (A. lappacea and A. harpophylla) or eight (A. scabra) or either five or eight or ten (T. australis) replicates. Treatments with the same letter do not differ significantly (P>0.05) from the control. ‘SDIW’ stands for ‘sterilized deionized water’. ‘nt’ stands for ‘not tested’. ‘2nd exp.’ refers to ‘second experiment’.
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Figure 4-1 Radicle elongation of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days of exposure to various concentrations of As(V) solutions. Arrows indicate emergent radicle.
Themeda australis’ RTI was not affected by As(V) up to 66.73 µM but it was significantly (P<0.05) reduced at 133.47 µM As(V) and was progressively reduced at higher concentrations compared to the control treatment (Table 4-3). This indicates that T. australis’ RTI toxicity threshold was between 66.73 and 133.47 µM As(V). 150
The RTI of A. scabra was considerably higher than the control at 13.34 µM As(V) and was not significantly different at any of the other As(V) concentrations tested, suggesting a RTI toxicity threshold >667.36 µM (Table 4-3).
For A. harpophylla, there was no significant (P>0.05) difference between RTI in any As(V) treated seeds compared to the control, also suggesting a RTI toxicity threshold >667.36 µM (Table 4-3).
4.4.2.2 Cu Copper at all concentrations tested did not affect significantly (P>0.05) the RTI of A. lappacea and T. australis compared to the control (Table 4-3), suggesting that the species’ RTI tolerance threshold to Cu was >200 µM.
4.4.2.3 Zn As with Cu, there were no significant (P>0.05) differences in the RTI of A. lappacea and T. australis in any treatment of Zn compared to the control (Table 4-3). This indicates that the species’ RTI toxicity thresholds were >500 µM.
4.4.2.4 Mn Manganese did not affect (P>0.05) RTI of A. lappacea, T. australis and A. harpophylla at any concentration tested compared to the control (Table 4-3), indicating a RTI toxicity threshold >10240 µM for these species.
4.4.2.5 Pb There was no significant difference between RTI of A. lappacea in Pb treatments compared to the control up to 4800 µM, where it was significantly reduced to 14.4% of the control value and further reduced to 0% in 9600 µM Pb (Table 4-3). At 9600 µM Pb, all the germinated seeds were without a radicle (shoot only) (Table 4-3 and Figure 4-2). RTI appeared to be a good indicator of Pb toxicity to A. lappacea during germination and the Pb effect was consistent across the treatments. The species’ RTI toxicity threshold to Pb was between 3840 and 4800 µM Pb.
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Figure 4-2 Radicle elongation of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days of exposure to various concentrations of Pb solutions. Arrows indicate emergent radicle.
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Themeda australis’ RTI was significantly (P<0.05) reduced compared to the control at Pb concentrations of 240 µM and higher (Table 4-3). This suggests that the species’ RTI toxicity threshold was <240 µM.
The RTI of A. scabra was considerably higher at 240 µM Pb compared to the control but remained not significantly (P>0.05) different from the control at all other Pb concentrations tested (Table 4- 3), indicating a RTI toxicity threshold >9600 µM for this species.
4.5 Discussion Our study suggests that radicle tolerance index (RTI) was the most sensitive indicator of the tolerance of native plant species to metal/loids during seed germination. The sensitivity order of tolerance indicators during germination was RTI > maximum germination percentage (Gmax) > shoot tolerance index (STI) > mean germination time (MGT). The finding in this study is in agreement with previous studies demonstrating high sensitivity of root length in response to toxicity of various metals (As, Cd, Cr, Pb) compared to other germination indicators such as germination rate, shoot length, maximum germination percentage and shoot biomass in both crop plants (Xiong, 1998, Jamali et al., 2006, He et al., 2008, Ahmad et al., 2011, Almas et al., 2012) and recognized metal-tolerant species (Rossato et al., 2011, Karimi et al., 2013). In addition, Hou et al. (2014) studied toxicity effects of chromium (Cr) on various plant species including Brassica oleracea, Cucumis sativus, Lactuca sativa, Triticum aestivum and Zea mays and found that root elongation was one of the most sensitive indicators for phytotoxicity of Cr in soil. Mishra and Choudhuri (1998) reported that shoot and root tolerance indexes were the most sensitive indicators for biomonitoring of phytotoxic effects of Pb and mercury (Hg) on rice cultivars. Roots are often used to measure heavy metal tolerance of plants because they are more responsive to metal toxicity in the environment (Xiong, 1998). This might be because roots function as the initial absorptive organs and are more directly exposed to metals compared to other plant tissues (Xiong, 1998, Hou et al., 2014). Therefore, the presence of metals may reduce the ability of roots to absorb water and nutrients from soils and may inhibit cell division in roots (Lefèvre et al., 2009, Hou et al., 2014); consequently, germinating seeds are not able to develop and grow. This is confirmed by the present study where substantial proportions of germinated seeds of A. lappacea were observed without a radicle in higher concentrations of As(V) and Pb treatments.
Overall, all native species were highly tolerant to the metal/loids tested during germination. However, species showed different metal/loid toxicity thresholds and levels of tolerance to As(V), Cu, Zn Mn and Pb using RTI as metal tolerance indicator. Acacia harpophylla and A. scabra were extremely tolerant to increasing concentrations of As(V), with no effects on the species’ RTI at the 153 highest concentration tested (667.36 µM), while A. lappacea was tolerant to slightly lower As(V) concentrations (between 400.41 and 533.88 µM). In contrast, T. australis was least tolerant to As(V), with its RTI being affected at relatively much lower As(V) concentrations (between 66.73 and 133.47 µM) than the other species. Overall, the As(V) tolerance rank during germination of the species was A. harpophylla = A. scabra > A. lappacea > T. australis. Although these species have been frequently used for rehabilitation of mine sites (Huxtable et al., 2005), their tolerance thresholds to As(V) had never been studied. This study was the first to establish the species tolerance threshold during germination in As(V) solutions. The four native species tested in this study were more tolerant or showed similar tolerance levels to As(V) during germination compared to previously studied plant species such as Phaseolus aureus, Helianthus annuus L., Hordeum vulgare L. and Triticum aestivum L. For example, radicle length of H. vulgare L. was reduced at 4 mg As (arsenate)/L (equivalent to 53.38 µM As) (Sanal et al., 2014), P. aureus at 10 µM As (arsenate) (Kaur et al., 2012), H. annuus L. at 6 - 10 mg As (sodium arsenate)/L (equivalent to 19.22 – 32.04 µM) (Imran et al., 2013), early root development of T. aestivum L. was reduced at 4 - 16 mg As (arsenite)/L (equivalent to 53.38 – 213.55 µM As) (Liu et al., 2005), and there was a gradual reduction of germination percentage, shoot and root length elongation of T. aestivum L. at 5 - 30 mg As (arsenate)/L (equivalent to 66.73 – 400.41 µM As) (Mahdieh et al., 2013).
The present study found that A. lappacea and T. australis were tolerant to high concentrations of Cu and both showed a similar level of tolerance to Cu, with RTI not affected by Cu at the highest concentration tested (200 µM). This suggests that both A. lappacea and T. australis have a greater tolerance to Cu during germination (>200 µM) than Alyssum montanum and Thlaspi ochroleucum for which germination percentage and growth rate of radicle were reduced at 80 and 160 µM Cu (Ouzounidou, 1995). Germination percentage of Zea mays was not affected at 12 ppm (equivalent to 188.82 µM Cu) but root and shoot lengths of the seedlings were greatly reduced at the same concentration of Cu (Mahmood et al., 2005). Root length of 6-day-old seedlings of Triticum aestivum L. was reduced at 50 ppm Cu (CuSO4.5H2O) (equivalent to 200.64 µM Cu) (Gang et al., 2013) and root and shoot length of Agropyron elongatum seedlings was reduced at 30 mg/L Cu (equivalent to 472.06 µM Cu) (Saberi and Shahriari, 2011). Other species showed higher tolerance to Cu, such as Pisum sativum, where seed germination percentage was reduced at 80 ppm Cu (equivalent to 1258.85 µM Cu) (Kunjam et al., 2015), maximum inhibition of seed germination percentage and decrease of shoot and root length of Cassia angustifolia Vahl occurred at 200 mg/L Cu (3147.12 µM Cu) (Nanda and Agrawal, 2016) and reductions of germination rate of Oryza sativa L. cv. Hwayeong occurred at 1 mM, shoot length at 0.2 mM and root formation at 250 µM (Ahsan et al., 2007).
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Our study showed that both A. lappacea and T. australis showed the same level of tolerance to Zn as RTIs of both species were not significantly affected by Zn and the RTI toxicity threshold was >500 µM. Rossato et al. (2011) reported RTI of A. lappacea was significantly reduced at 10 mM Zn in solution during germination. The tolerance of T. australis to toxic concentrations of Zn during germination was not previously known. Tolerance of other Australian native species to Zn has been studied by Reichman et al. (2001) using much lower concentrations (up to 100 µM Zn) than those tested in the study (up to 500 µM), but during the seedling growth stage. Kunjam et al. (2015) reported Zn at 80 ppm (equivalent to 1223.61 µM Zn) did not affect root growth of P. sativum seedlings during germination and growth. Zinc at 50 mg/L (equivalent to 764.75 µM Zn) also markedly reduced root and shoot lengths of 28-day-old Cassia angustifolia seedlings (Nanda and Agrawal, 2016), and root growth of Salvia coccinea was inhibited at 100 mg/L (equivalent to 1529.51 µM Zn) (Stratu et al., 2014). Dorycnium pentaphyllum germination was slightly delayed and final germination percentage was greatly reduced at 10000 µM Zn (Lefèvre et al., 2009).
Astrebla lappacea, T. australis and A. harpophylla were equally tolerant to Mn at all concentrations tested with a RTI toxicity threshold >10240 µM. The species tolerance rank to toxic concentrations of Mn during germination was A. lappacea = A. harpophylla = T. australis. Manganese toxicity thresholds for these species have not been studied previously. Other species, such as Nicotiana tabacum, have been found to tolerate 2 mM Mn during tissue regeneration as callus induction, callus growth and shoot regeneration were not affected at 2 mM Mn (Santandrea et al., 1997). In Arabidopsis, primary root growth was only affected by 2 mM Mn (Zhao et al., 2017). In addition, germination of Xanthium sibiricum was only affected by Mn concentrations above 5000 μM where germination potential, germination index, vigor index, root length, fresh weight, dry weight and root-shoot ratio were greatly reduced (Pan et al., 2017).
Austrostipa scabra, A. lappacea and T. australis were tolerant to high concentrations of Pb. Based on RTI as a tolerance indicator, the order of species tolerance to Pb was A. scabra (>9600 µM) > A. lappacea (between 3840 and 4800 µM) > T. australis (<240 µM). Toxicity of Pb in solutions to germination of A. lappacea was studied previously by Rossato et al. (2011), who reported a significant decrease in RTI (0% of the control value) at 9600 µM Pb. In contrast, tolerances of germinating A. scabra and T. australis seeds to Pb have never been tested. In comparison, previous studies reported that the germination percentage of Matricaria chamomilla was not affected by Pb at 60 µM but root and shoot lengths and weight of seedlings of the species were reduced at 30 µM Pb (Saderi and Zarinkamar, 2012). Germination rate of Triticum aestivum L. was reduced at 0.3 mM Pb but root and leaf elongation of germinated seeds were reduced at 1.5 mM Pb (Lamhamdi et
155 al., 2011), and in Festuca arundinacea inhibitory effects of Pb on relative germination rate and germination index were apparent at 1000 mg Pb/L (equivalent to 4826.25 µM) (Lou et al., 2017).
4.6 Conclusion Radicle tolerance index was found to be the most sensitive indicator of As(V), Cu, Zn, Mn and Pb tolerance in A. lappacea, T. australis, A. scabra and A. harpophylla during seed germination. All native species were highly tolerant to the metal/loids tested, however, they showed different metal toxicity thresholds and levels of tolerance based on RTI as a metal tolerance indicator during germination. Our study was only able to identify toxic concentration ranges for some metal/loids and for some of the species, and further work will be needed to identify more precise RTI toxicity thresholds (actual concentrations or at least limited concentration ranges as opposed to the ranges identified here). All four species could be classified as metallophytes, confirming their current suitability and use for mine site rehabilitation. This work may also serve as a basis for future studies on metal/loid tolerance of other plant species during germination. The resulting RTI toxicity thresholds will only apply to germinating seeds and additional studies will be required to confirm whether these thresholds can also be applied to species during plant establishment.
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Chapter 5. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings
The chapter was published as a research article in a high ranking refereed journal, Journal of Hazardous Materials, 2013 and in international conference proceedings with the following details:
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2013. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248– 249, p.442– 450.
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2012. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia, page 517-532.
Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Schmidt, S. 2011. Micron-sized metal-binding hydrogel particles improve germination and root elongation of Australian metallophyte grasses (Astrebla lappacea and Austrostipa scabra) on highly polluted waste rock and tailings. Proceedings of the 32 Australasian Polymer Symposium 13-17 February 2011. Coffs Harbour, New South Wales, Australia.
For uniformity of the thesis, some changes in the formatting have been made including omitting the ‘article highlights’ and inclusion of the references in the complete thesis reference list.
The idea and design of the study were generated by the candidate and Drs Laurence Rossato, Alex Pudmenzky, David Doley, Michael Whittaker and Prof Susanne Schmidt. Production of metal- binding particles (X3) was funded by the University of New South Wales Centre for Advanced Macromolecular Design (CAMD). X3 particles were prepared by Dr Michael Whittaker and the candidate, with assistance from Dr Laurence Rossato and Dr Alex Pudmenzky. The experiment was conducted by the candidate with the assistance of Mr Craig Claremont (a 2010 UQ Summer Scholarship Student) and advice from Drs Laurence Rossato, David Doley and Alex Pudmenzky. Photographs taken during the experiment were by Drs Alex Pudmenzky and Laurence Rossato. 157
Data analysis was conducted by the candidate with assistance from Drs Laurence Rossato, David Doley, Alex Pudmenzky and Michael Whittaker. Tailings and waste rock substrates were donated by operating and closed mine sites in Queensland (Australia).
The manuscript was written by the candidate with editorial advice from Drs Laurence Rossato, David Doley, Alex Pudmenzky and Michael Whittaker.
5.1 Abstract Metal contamination of landscapes as a result of mining and other industrial activities is a pervasive problem worldwide. Metal contaminated soils often lack effective vegetation cover and are prone to contaminant leaching and dispersion through erosion, leading to contamination of the environment. Metal-binding hydrogel particle amendments could ameliorate mine wastes prior to planting and enhance seedling emergence. In this study, micron-size thiol functional cross-linked acrylamide polymer hydrogel particles (X3) were synthesised and tested in laboratory-scale experiments on phytotoxic mine wastes to determine their capacity to: (i) increase substrate water holding capacity (WHC); (ii) reduce metal availability to plants to below the phytotoxicity threshold; and (iii) enhance germination characteristics and early radicle development of two Australian metallophyte grasses under limiting and non-limiting water conditions. Addition of X3 to mine wastes significantly increased their WHC and lowered toxic soluble metal concentrations in mine waste leachates. Germination percentages and radicle elongation of both grasses in wastes were significantly increased. Highest germination percentages and greater radicle development were recorded in X3 amended wastes under water limited conditions, suggesting that X3 was able to ameliorate metal toxicity to radicles and provide moisture, which improved the imbibition and consequent germination of the seeds.
Keywords: mine wastes, metal-binding hydrogel particles, metallophytes, soil water holding capacity, remediation
5.2 Introduction Heavy metal contaminated soils resulting from mining and other industrial activities are major environmental problems because they create unfavourable conditions for plant growth due to the lack of soil structure, low nutrient concentration, high salinity, extreme soil pH and elevated concentrations of toxic elements such as aluminium (Al), copper (Cu), lead (Pb), nickel (Ni), zinc (Zn) and arsenic (As) (de Varennes et al., 2011). As a consequence, contaminated soils, tailings and waste rock may lack protective layers of vegetation that prevent dispersion of toxic contaminants to adjacent aquifers and other areas (Alvarez et al., 2003, Antonijevic et al., 2012, Mendez and Maier,
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2008a). The contaminants may be leached, absorbed by vegetation or retained by the soil (Alvarez et al., 2003) so they may enter the food chain and put human and wildlife health at risk.
A vegetative cover on mine wastes reduces dispersion of toxic elements to surrounding ecosystems (de Varennes et al., 2011). Phytoremediation is the use of metal tolerant plants (i.e. metallophytes) to remove or stabilise toxic elements in soils and is a cost-effective and ecologically advantageous approach compared to other remediation techniques such as excavation and disposal to landfill (Hegazy et al., 2011, Nedunuri et al., 2010, Salt et al., 1995). Native plant species are often preferred for phytoremediation as they are likely to be resilient to local stresses such as low soil nutrient and organic matter contents and climate (for instance drought in arid and semiarid environments) (Mendez and Maier, 2008a). Moreover, they are generally not invasive, and thereby maintain and possibly increase local plant biodiversity (Keeling and Werren, 2005, Mendez and Maier, 2008a). Grasses can often achieve rapid ground cover and limit short term erosion (Williams and Currey, 2002) and shrubs and trees offer a broad canopy cover and establish deep root systems that prevent erosion over the long term (Belsky et al., 1989, Tiedemann and Klemmedson, 2004). However, successful plant establishment of even metallophytes may be limited or impossible when metal contamination in soils exceeds plant toxicity thresholds. So, enhancing the quality of heavily impacted soils prior to revegetation is often a prerequisite for effective phytoremediation.
Nanotechnology offers a further approach to in situ remediation of metal contaminated soils which could complement existing techniques (Sun et al., 2006). The use of nanoparticles (surface- functionalized polymer and nanoscale zero-valent iron) to remove metals such as mercury (Hg), Pb, cadmium (Cd) and chromium (Cr) from contaminated water and soils has been reported previously (Bell et al., 2006, Singh et al., 2012). Plant establishment in polluted sites could be enhanced through the targeted design and application of micron- or nano-sized particles that have the capacity to bind toxic soluble metals in soils and improve plant water relations (Rossato et al., 2011). For example, application of insoluble hydrophilic polyacrylate polymers in mine soil contaminated with Pb2+ has been shown to increase soil water holding capacity (Guiwei et al., 2008), reduce bioavailability of toxic trace elements such as Pb2+ and Cu2+, and enhance plant growth (de Varennes and Queda, 2005, Guiwei et al., 2008, Torres and de Varennes, 1998). Application of hydrophilic polymers from diapers at 0.3% (m/m) to mine soil has also been found to induce faster establishment of plant cover (Qu and de Varennes, 2010). Nano-sized vivianite was effective in reducing the leachability and bioaccessibility of Cu(II) in calcareous, neutral and acidic soils (United States-E.P.A, 2005). Rossato et al. (2011) showed that micron-size thiol functional cross- linked acrylamide polymer particles (called X3) were able to reduce the available solution
159 concentrations of Pb2+ (9.65 mM), Cu2+ (4 mM) and Zn2+ (10 mM) by 86.5, 75.5 and 63.84%, respectively. They also reported that X3 was not toxic to seed germination as it allowed normal germination and root elongation of the native metallophyte curly Mitchell grass (Astrebla lappacea) at phytotoxic available concentrations of Pb2+ (9.65 mM) and Zn2+ (10 mM).
The objectives of this study were (i) to assess the potential of X3 to bind toxic soluble metals in extremely saline and metal contaminated mine waste rock and tailings, and to increase substrate water holding capacity (WHC), and (ii) to investigate X3 effect on the germination characteristics and early radicle development of two Australian metallophyte grasses, Astrebla lappacea and Austrostipa scabra, under water limiting and non-limiting conditions.
5.3 Materials and methods 5.3.1 Substrate collection and characterisation Substrates were collected from two mine sites where revegetation attempts had been unsuccessful for over 30 years. Waste rock (top 10 cm only) was collected from an abandoned gold mine in Queensland. Tailings (representative sample from the top 7 m) were collected from an active base metal mine in Queensland, Australia. River sand with no metal contamination was used as a control and was collected from Mt. Tamborine (Australia) and acid washed and thoroughly rinsed with deionised water (DIW) prior to experiments.
All substrates were air dried, crushed, mixed and sieved (<2 mm) to produce a fine, homogenous medium. Substrates were then analysed for pH and EC in a soil water suspension (4 g of soil to 20 mL distilled water) after shaking for 1 h by inversion at ∼40 rpm in a Heidolph ReAx shaker and standing for 1 h (Rayment and Higginson, 1992). Plant available metals were extracted using 0.01
M CaCl2 (4 g of soil to 40 mL 0.01 M CaCl2) (Menzies et al., 2007). Determination of total metals was achieved by reducing the substrate to fine particles using a ball mill (Planetary Ball Mills PM 200, RETSCH, Germany). Acids (5 mL nitric acid 70% from Labscan Asia Co., Ltd., Bangkok, 2 mL hydrochloric acid 32% and 2.5 mL hydrofluoric acid 50% from Ajax Finechem, Australia) were added to 150–200 mg of sample which was then heated in a microwave digester (MDS-200, CEM Corporation) at 120 atm (12.16 MPa) and 185 ◦C for 30 min.
Samples were then analysed for a range of metals using inductively coupled plasma optical emission spectroscopy (ICP-OES) or inductively coupled plasma mass spectrometry (ICP-MS). A National Institute Standard and Technology (NIST) Standard Reference Material (SRM) (2709 San Joaquin soil), which has certified values for most of the metals, was used to verify the measurements.
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5.3.2 Determination of metal-binding and water holding capacities of X3 in mine substrates under various pH conditions Micron-size thiol-functional cross-linked acrylamide particles (X3) were synthesised as described in Rossato et al. (2011). Mixing experiments were carried out to test the efficiency of X3 in reducing available concentrations of metals and increasing the WHC of the waste rock and tailings over a large pH range (6.3–2) in DIW and 0.01 M CaCl2. Two different percentages of X3 were added to each substrate based on their plant available metal concentrations so as to bind all or half of the soluble metals as described by Rossato et al. (2011). For DIW treatments (pH 6.3), percentages of X3 added to substrates were 18.4 and 9.2% DW for waste rock and 3.2 and 1.6% DW for tailings.
For 0.01 M CaCl2 treatments (pH 6.3–2), 9.2% DW and 1.6% DW of X3 were added to waste rock and tailings, respectively. Required amounts of X3 were thoroughly mixed with substrates for 12 h using a shaker (Gerhardt Rotoshake RS12 Elution Shaker) at 5 rpm.
Ten mL of DIW (pH 6.3) or 0.01 M CaCl2 at various pH were added to one gram of waste rock or tailings unamended or amended with X3 in 50 mL polypropylene conical tubes. Samples were shaken overnight (12 h) at 5 rpm and then centrifuged (3000 rpm for 30 min at 22◦C). The supernatant (mine substrate leachate), which contained the fraction of free soluble metals not bound to the particles, was pipetted into 10 mL PPTR tubes and 25 μL of nitric acid (70%) was added per 1 mL of supernatant with stirring to digest the samples prior to metal analysis via ICP-OES/MS. Three replicates per treatment were used.
The metal-binding efficiencies of the particles in mine substrates were calculated using the formula by Rossato et al. (2011):
(1)
The water holding capacities (WHC) of waste rock and tailings under saturated conditions were calculated using the formula below (Rossato et al., 2011):
(2)
where Wh is the hydrated weight of unamended or X3 amended waste rock or tailings and Wd is dry weight of unamended or X3 amended waste rock or tailings.
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5.3.3 Measurement of germination characteristics and radicle elongation of A. lappacea and A. scabra in mine substrates amended with X3 under different water regimes Viable seeds of A. lappacea and A. scabra were obtained from Native Seeds (New South Wales, Australia). They were not dormant as the seeds had been collected and stored dry at 3–5 ◦C for more than 12 months prior to the germination experiments.
The amounts of X3 particles applied to waste rock and tailings were calculated so as to bind all toxic soluble metal as described by Rossato et al. (2011) and were 18.4 and 3.2% DW for waste rock and tailings, respectively. Particles were mixed with the waste rock and tailings overnight using a shaker (Gerhardt Rotoshake RS12 Elution Shaker) at 5 rpm. Uncontaminated acid washed river sand was used for the control.
Seeds were surface-sterilised in 20% sodium hypochlorite for 10 min and then rinsed three times in sterilised deionised water (SDIW) for one min. Thirteen grams of waste rock or tailings or 20 g of sand (control) with or without particles were placed in each Petri dish (disposable dish 90 mm × 25 mm) and a predetermined volume of SDIW was added to each dish to bring it to field capacity. Forty seeds of each species were placed on top of the substrates in each Petri dish, with 5 replicates for each treatment. Two different water regimes (field capacity and water-limited) were tested to investigate the potential of the X3 particles to supply water during a normal germination period.
Petri dish treatments maintained at field capacity were sealed with parafilm and placed in a transparent plastic zip resealable bag to minimise water loss. In water-limited treatments, substrates were watered to field capacity on the first day of the experiment but received no additional water thereafter. Water-limited Petri dishes were not sealed with parafilm nor placed in plastic bags. All the Petri dishes were placed in a germination cabinet under controlled conditions (30˚C day, 25˚C night, 12 h light per day).
Germinated seeds were counted and removed daily, within a laminar flow cabinet until the maximum germination percentage was reached, i.e. for 15 and 21 days for A. scabra and A. lappacea, respectively. Seeds were considered germinated when 0.5 mm of radicle had emerged.
Mean germination times (MGT) were calculated using the following formula of Ellis and Roberts (1981):
(3)
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where D is the number of days counted since the beginning of the experiment and n is the number of germinated seeds on day D.
Radicle length of germinated seeds was measured from the radicle-shoot junction to the tip of the longest radicle after 3 days for A. scabra and 4 days for A. lappacea.
The radicle tolerance index (RTI) was used as sensitive indicator of the alleviation of substrate toxicity by X3 addition (Rossato et al., 2011) and was calculated using the following formula (Ellis and Roberts, 1981):
(4) where control substrate is sand treatment alone or with X3 particles as appropriate.
5.3.4 Determination of pH, plant available metal concentrations and cumulative evaporation in germination substrates Substrates used for germination experiments were collected and air dried at the end of the experiment for pH and plant available metal analysis following the methods described in Section 5.3.1.
Cumulative evaporation in germination substrates was measured by weighing germination Petri dishes on every second day until Day 14 for A. scabra and Day 20 for A. lappacea. For field capacity treatment, weighing of the Petri dishes was done before addition of SDIW.
5.4 Data analysis Data were presented as means and standard errors of either three (mixing experiment) or five (germination experiment) replicates, and they were tested for significant differences between treatment means using Student’s t-tests in Microsoft Office Excel 2007. Analyses were conducted using a 95% confidence interval and statistical significance was indicated by P-values P<0.05.
5.5 Results 5.5.1 Characteristics of substrates Characterisations of the waste rock and tailings are presented in Table 5-1. Aluminium and copper were the major contaminants of waste rock and total concentrations of these metals (Al = 29,000 mg kg-1 and Cu = 1,389 mg kg-1) reached recognised soil plant toxicity concentrations (Kopittke et al., 2010, Mendez and Maier, 2008a). The plant available concentrations of Al (308 mg kg-1) and Cu (34.6 mg kg-1) were also high compared to toxicity threshold limits for plants of 0.5 mg kg-1 of
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Al or Cu (Mulvey and Elliott, 2000). Other metals (Mn and Zn) were also present in the waste rock although their plant available concentrations were lower than plant toxicity concentrations for these metals. Concentrations of total Na and S in waste rock exceeded plant toxicity limits (Mulvey and Elliott, 2000). Unlike waste rock, the main contaminants in tailings were Mn and Zn with total and plant available concentrations of these metals exceeding plant toxicity limits (Kopittke et al., 2010, Mendez and Maier, 2008a). Total Al and Cu were also high but the bioavailabilities of these metals were lower than the plant toxicity concentrations. The tailings also contained total Na and S with concentrations above toxicity limits for plants (Mulvey and Elliott, 2000). Sand showed no trace of metal contamination (Table 5-1) and was an appropriate control for the germination experiment.
The waste rock was also very acidic (pH 3.03±0.003). With very low pH, the plant availabilities of metals in waste rock, such as Al, Cu and Zn are likely to increase as these metals are more available at low pH. In contrast, pH of the tailings was neutral (7.09±0.006) which would reduce metal availability to plants. Both substrates were extremely saline (EC = 2.79±0.003 and 3.63±0.006 dS.m-1 in waste rock and tailings, respectively (The State of Queensland, 2011). These ECs are toxic to plants, reducing growth, early maturation and plant yield (50% yield reduction was observed in grasses) (Mulvey and Elliott, 2000).
Table 5-1 Characterisation of waste rock, tailings and sand (control). Results are given as the mean ± standard error (SE) (n=3).
Elements Waste rock Tailings Sand (control) Total (mg kg-1) Plant available Total (mg kg-1) Plant available Total (mg kg-1) Plant available (0.01 M CaCl2) (0.01 M CaCl2) (0.01 M CaCl2) (mg kg-1) (mg kg-1) (mg kg-1)
Al 29,000±703 308±4.5 14,566±388 0.9±0.4 15,049±294 0.0±0.0
Cu 1389±11 34.6±0.6 842±16 0.2±0.0 0.0±0.0 0.0±0.0
Na 8279±109 - 602±25 - 4051±97 -
S 23,856±454 - 44,070±792 - 193±14 -
Mn 452±52 8.0±0.2 2706±51 76.7±1.7 166±1.6 0.2±0.0 4.6±0.1 3647±75 18.5±0.4 8±0.7 0.0±0.0 Zn 368±6.4
5.5.2 Effect of X3 particles on soluble metal concentrations in mine waste leachates X3 particles significantly reduced soluble concentrations of Al and Cu in leachate of amended waste rock (Table 5-2). Concentrations of soluble metals in waste rock amended with 9.2% DW of X3 were decreased by about 23% for Al (from 26.5 to 20.3 mg L-1) and 27% for Cu (from 3.6 to 2.7 mg L-1) compared to unamended waste rock. The concentrations of Al and Cu in amended waste
164 rock were further reduced by 40% (to 12 mg L-1) and 43% (to 1.5 mg L-1) respectively, when the percentage of X3 was increased to 18.4%. pH did not affect the functionality of X3 in reducing soluble concentrations of Al and Cu in leachate of X3-amended waste rock (9.2% DW) between pH 6.3 and 3.2 (Table 5-2). The average reduction of Al and Cu soluble concentrations was almost constant at 23.5±1.5% and 26.7±1.6% between pH 6.3 and 3.2. Below pH 3.2, percentage metal reduction decreased to 5.6±1.0% for Al and 0% for Cu, which suggests a loss of particle functionality in extremely acidic conditions.
Table 5-2 Soluble concentrations of metals in leachate of waste rock at different percentages of X3 and under various pH conditions in DIW and 0.01 M CaCl2. Data are presented as the mean ± standard error (SE) (n=3); nt stands for “not tested”.
Extractant pH Solution metal concentration (mg L-1) in waste rock leachates Al Cu Percentage of X3 to waste rock (% DW) Percentage of X3 to waste rock (% DW) 0 9.2 18.4 0 9.2 18.4 DIW 6.3 26.5±0 20.3±0.6 12±0.1 3.6±0 2.7±0.1 1.5±0
0.01 M CaCl2 6.3 31±0.3 21.6±2 nt 3.7±0 2.4±0.2 nt
0.01 M CaCl2 5.2 30.4±0.8 23.5±0.1 nt 3.6±0.1 2.7±0.1 nt
0.01 M CaCl2 4.5 31.2±0.2 23.6±0.8 nt 3.7±0 2.7±0.1 nt
0.01 M CaCl2 3.2 33.3±0.7 25.9±0.3 nt 4±0.1 3±0 nt 0.01 M CaCl2 2 47±0.8 45.3±0.3 nt 5±0.1 5.1±0 nt
As compared with waste rock, tailings had relatively low soluble metal concentrations in leachates and addition of X3 did not significantly (P>0.05) decrease concentrations of soluble metals (Data not shown).
5.5.3 Water holding capacities of mine substrates amended with X3 WHCs of unamended and X3 amended waste rock and tailings were tested to examine the potential of X3 to provide water for imbibition and germination of seeds. X3 particles greatly improved WHCs of both waste rock (at 9.2 and 18.4% DW of X3) and tailings (at 1.6 and 3.2% DW of X3) (Table 5-3). Addition of 9.2% of X3 to waste rock and 1.6% of X3 to tailings increased WHCs of the substrates approximately 4.5-fold (42 - 190%) and 2-fold (29 - 60%), respectively. As percentages of X3 increased to 18.4% for waste rock and 3.2% for tailings, WHCs of the substrates increased significantly, to around 266 and 79%, respectively.
There was no effect of pH on the capacity of X3 particles to increase WHCs of waste rock and tailings (Table 5-3). Compared with unamended waste rock and tailings, WHCs of X3-amended 165 waste rock (9.2% DW of X3) and tailings (1.6% DW of X3) in 0.01 M CaCl2 solution increased by nearly 5 times from 44 to 210% and by 2 times from 29 to 59%, respectively, and were constant over the range of pH tested (6.3 to 2).
Table 5-3 Water holding capacities of waste rock and tailings amended with different percentages of X3 and under various pH conditions in DIW and 0.01 M CaCl2. Data are presented as the mean ± standard error (n=3); nt stands for “not tested”. Treatment Extractant pH Substrate water holding capacities (% DW) Percentage of X3 to substrates (% DW)
0% 9.20% 18.40% Waste rock DIW 6.3 42.2±0.3 189.8±11 265.8±4.9
0.01 M CaCl2 6.3 36.3±1.3 211.9±18.9 nt
0.01 M CaCl2 5.2 43.4±3.7 201.9±14.1 nt
0.01 M CaCl2 4.5 47.4±1 199.3±13.6 nt
0.01 M CaCl2 3.2 47.1±1 187.7±5.6 nt
0.01 M CaCl2 2 43.7±2.1 243.0±26.7 nt Treatment Extractant pH Substrate water holding capacity (% DW) Percentage of X3 to substrates (% DW)
0% 1.60% 3.20% Tailings DIW 6.3 28.8±0.3 59.6±5.3 78.8±4.4
0.01 M CaCl2 6.3 28.5±1.2 58.3±1.3 nt
0.01 M CaCl2 5.2 33.0±1.7 62.2±1.7 nt
0.01 M CaCl2 4.5 29.6±1.1 59.7±1.4 nt
0.01 M CaCl2 3.2 27.6±0.8 59.4±0.2 nt
0.01 M CaCl2 2 27.8±0.4 57.0±3.1 nt
5.5.4 Effect of X3 on seed germination 5.5.4.1 Effect of X3 on pH, plant available metal concentrations and cumulative evaporation of germination substrates Waste rock used for the germination experiment was analysed for pH and plant available metal concentrations after the experiment concluded (Table 5-4). Regardless of the water treatment and the grass species, the addition of X3 increased the pH significantly (P<0.05), from around 3.2 to up to 3.9.
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Table 5-4 pH and plant available metal concentrations (0.01 M CaCl2) in unamended and X3 amended (18.4% DW) waste rock used for the germination of A. lappacea and A. scabra under two different water regimes. Data are presented as the mean ± standard error (SE) (n=3). Treatment pH Plant available metal concentration (mg kg-1)
Al Cu a) A. lappacea Waste rock + water limited 3.3±0 239±1 27.6±0.2 Waste rock + field capacity 3.3±0 212.6±3.4 23.1±0.4 Waste rock + X3 + water limited 3.9±0 38.2±0.8 6.0±0.3 Waste rock + X3 + field capacity 3.8±0 46.3±1 6.8±0.2 b) A. scabra
Waste rock + water limited 3.3±0 281.2±2.3 31.8±0.2 Waste rock + field capacity 3.2±0 290±3.4 32.2±0.3 Waste rock + X3 + water limited 3.9±0 42.6±2.1 6.4±0.2 Waste rock + X3 + field capacity 3.9±0 54.0±1.6 7.3±0.1
Plant available concentrations of Al and Cu in waste rock decreased remarkably and significantly (P<0.05) with the addition of X3 in the substrates of both grass species (Table 5-4). Metal concentrations were lowered by between 78 and 85% and between 70 and 81% for water limited and field capacity treatments, respectively, and maintained below the phytotoxicity thresholds for A. lappacea and A. scabra.
In tailings, regardless of the water treatment and the grass species, addition of X3 (3.2% DW) did not modify pH and plant available metal concentrations, which were not significantly different (P>0.05) from those in unamended tailings (Data not shown).
Cumulative evaporation from the substrates for both species was also measured during the germination experiment (Figure 5-1). Unamended waste rock and tailings maintained uniform evaporation for 10 days, by which time they appeared dry. For unamended sand (control), evaporation ceased after 14 days.
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Figure 5-1 Cumulative water loss from sand, waste rock and tailings for A. lappacea (a and b) and A. scabra (c and d) in Petri dishes maintained at field capacity or water limited conditions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean (n=5).
In the presence of X3, cumulative evaporation increased linearly for at least 20 days in waste rock and control sand (Figure 5-1a) and for 18 days in tailings under restricted water regime (Figure 5- 1b) for A. lappacea. This shows that X3 increased the capacities of both substrates to hold water and release it progressively during the germination period under restricted water conditions. The cumulative water loss from waste rock was greater than from tailings as the percentage of X3 added 168 to waste rock was 18.4% DW against only 3.2% DW for tailings (Figure 5-1a and b). In field capacity treatments, evaporation rates were low in both substrates amended with X3 because the Petri dishes were placed in plastic zip resealable bags to reduce water loss (Figure 1a and 1b). Similar trends were observed for A. scabra (Figure 5-1c and d).
5.5.4.2 Effect of X3 on germination percentages and mean germination time of A. lappacea and A. scabra in germination substrates The capacity of X3 particles to bind metals and provide water to plants in contaminated soils was tested in a germination experiment using Australian native metallophyte grasses, A. lappacea and A. scabra. X3 particles were not toxic to seed germination and significantly (P<0.05) enhanced germination percentages of A. lappacea and A. scabra in both waste rock and mine tailings (Table 5-5). Increases in germination percentages under restricted water regime conditions were significantly (P<0.05) higher than in field capacity treatments except for A. lappacea on tailings.
In A. lappacea, maximum germination percentages at field capacity increased significantly (P<0.05) with X3 amendment, from 0.5 to 29.6% and 12.9 to 23% for waste rock and tailings, respectively. The increase was higher for waste rock under the water limited regime (43.5% in X3 amended compared to 9.4% in unamended waste rock). The maximum germination percentage of A. lappacea in tailings was about 10.4% in the water limited treatment versus only 0.9% in the unamended water limited control. In A. scabra, maximum germination percentages increased significantly (P<0.05) on X3 amended substrates (from 0 to 12% and 0.5 to 24.5% in waste rock; from 8 to 13.5% and 6.5 to 21.5% in tailings) under field capacity and water limited regimes, respectively. Overall, maximum germination percentages of both species in X3 amended waste rock and tailings at field capacity were not significantly different (P>0.05) from the relevant sand controls. With the exception of A. scabra in amended waste rock (which was significantly higher, P<0.05) and in amended tailings (which was not significantly different, P>0.05), maximum germination percentages under water restricted conditions were lower (P<0.05) than the amended sand controls but significantly higher (P<0.05) than the unamended waste controls.
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Table 5-5 Maximum germination percentages (MaxGerm %) and mean germination time (MGT) of A. lappacea and A. scabra on unamended and X3 amended waste rock (18.4% DW of X3) and tailings (3.2% DW of X3) under different water regimes. Results are given as the mean ± standard error (SE) (n=5). Within each substrate, treatments with the same letter do not differ significantly (P>0.05); asterisks indicate values based on one replicate only.
Treatment A. lappacea______A. scabra______MaxGerm MGT MaxGerm MGT
(%) (Days) (%) (Days) (a) Waste rock Sand + water limited 20.5±2.9a 5.3±0.3a 15.5±3.0ac 3.1±0.1a Sand + field capacity 29.7±6.5aef 4.1±0.1b 18±2.4ac 3.0±0.2a Sand + X3 + water limited 78.4±3.9b 5.4±0.3a 14.0±3.0a 5.2±0.8b Sand + X3 + field capacity 18.4±6.5acf 4.5±1.1ab 15.0±3.5ac 3.1±0.4a Waste rock + water limited 9.4±3.2c 4.6±0.3ab 0.5±0.5b 6* Waste rock + field capacity 0.5±0.5d 4* 0.0±0.0b - Waste rock + X3 + water limited 43.5±4.0e 4.7±0.4ab 24.5±2.7c 4.7±0.4b Waste rock + X3 + field capacity 29.6±3.0f 5.2±0.3a 12.0±2.0a 3.9±0.7ab
(b) Tailings Sand + water limited 20.5±2.9ade 5.3±0.3a 20.5±3.0ac 2.7±0.1a Sand + field capacity 29.7±6.5a 4.1±0.1b 13.5±0.6bd 2.9±0.2a Sand + X3 + water limited 80.1±2.9b 4.1±0.1b 22.0±2.3a 5.1±0.6b Sand + X3 + field capacity 18.4±1.9ade 5.3±0.7abd 13.5±1.7cd 2.5±0.2a Tailings + water limited 0.9±0.9c 4* 6.5±1.3e 4.4±0.4b Tailings + field capacity 12.9±2.4de 11.5±1c 8.0±1.5e 7.5±1.8b Tailings + X3 + water limited 10.4±3.6e 3.4±0.9ab 21.5±4.9ad 5.0±0.3b Tailings + X3 + field capacity 23±2.4a 6.7±0.3d 13.5±1.5cd 5.0±0.3b
Addition of X3 particles to waste rock and tailings did not significantly (P>0.05) affect mean germination time (MGT) of either A. lappacea or A. scabra, with the exception of A. lappacea in tailings at field capacity, where germination was accelerated by almost five days in the presence of X3 (Table 5-5). MGTs of both grasses in both substrates was not affected by addition of X3 to limited water treatments and were not significantly different (P>0.05) from those at field capacity, with the exception of A. lappacea, which germinated nearly 3.5 days faster in tailings under water limited conditions. Regardless of water treatment, MGTs of both species in X3 amended mine substrates were not significantly different (P>0.05) from the appropriate sand control, except for A.
170 scabra in amended tailings under the limited water regime, which was slightly higher (P<0.05) than the amended sand control.
5.5.4.3 Effect of X3 particles on radicle elongation of A. lappacea and A. scabra in mine substrates X3 significantly enhanced radicle elongation of both species in contaminated waste rock and tailings (Figure 5-2 and Table 5-6). In general, radicle elongation of A. lappacea was greater under water limited conditions but the opposite was found for tailings. In A. scabra, the longest radicles were observed in amended waste rock at field capacity and in amended tailings under water limited conditions. Under the restricted water regime, radicle length of A. lappacea was up to 19 and 13.5 mm in amended waste rock and tailings (respectively) as compared with 2 and 3 mm in unamended treatments. In A. scabra, radicle length was up to 5.5 and 4.5 mm in amended waste rock and tailings (respectively) versus 0 and 1 mm in unamended treatments. At field capacity, radicle length of A. lappacea reached 5 and 17.5 mm in amended waste rock and tailings (respectively) compared to 0.5 and 0.0 mm in unamended treatments. Radicle length of A. scabra reached 7 and 3.5 mm in waste rock and tailings (respectively) as compared to 0.0 and 1.0 mm in unamended treatments.
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Figure 5-2 Radicle elongation of A. lappacea after 4 days and A. scabra after 3 days at field capacity and in water limited regime on unamended and X3 amended waste rock (18.4% DW) and tailings (3.2% DW). Arrows indicate emergent radicle.
Radicle tolerance index (RTI) of A. lappacea and A. scabra was also greatly increased by the addition of X3 in both waste rock and tailings (Table 5-6). In the presence of X3, the highest RTI of A. lappacea was recorded in tailings at field capacity (138%) which was significantly (P<0.05) higher than the sand control (100%), followed by waste rock under the water limited regime (108%), which was similar (P>0.05) to the sand control (100%). Although the RTIs of the species in X3 amended waste rock at field capacity (almost 53%) and tailings under water limited condition (46%) were lower than in control sand (100%), they were significantly (P<0.05) higher than in unamended waste rock (1.1%) and tailings (13.6%) under the same water regime. Regardless of water supply treatment, the RTI of A. scabra in waste rock and tailings in the presence of X3 was not significantly different (P>0.05) from the sand control. Large standard errors in some of the treatments were due to variable radicle extension. 172
Table 5-6 Radicle tolerance index (RTI) of A. lappacea and A. scabra in unamended and X3 amended waste rock (18.4% DW of X3) and tailings (3.2% DW of X3). Results are given as the mean ± standard error (SE) (n=5). Within each substrate, treatments followed by the same letter do not differ significantly (P>0.05); asterisks indicate values based on one replicate only.
Treatment RTI (%) A. lappacea A. scabra Control = sand Control = sand + X3 Control = sand Control = sand + X3 (a) Waste rock, water limited
Sand 100±0.0a - 100±0.0a - Sand + X3 222±80.3a 100±0.0a 67.7±27.3a 100±0.0a Waste rock 32.3±22.9b - 0±0b - Waste rock + X3 - 108±29.5a - 65.5±40.6a
(b) Waste rock, field capacity Sand 100±0.0a - 100±0.0a - Sand + X3 33±21a 100±0.0a 34.8±22.3b 100±0.0a Waste rock 1.1±1.1b* - 0±0b - Waste rock + X3 - 52.8±2.8b - 85±50.6a
(c) Tailings, water limited Sand 100±0.0a - 100±0.0a - Sand + X3 324±124.7a 100±0.0a 150±38a 100 ± 0.0a Tailings 13.6±13.6b* - 24.7±10.4b - Tailings + X3 - 46.1 ± 18.7b - 74.5±24.8a
(d) Tailings, field capacity Sand 100±0.0a - 100±0.0a - Sand + X3 34.2±19.7b 100±0.0a 129±29.3a 100±0.0a Tailings 0±0c - 6.7±6.7b - Tailings + X3 - b - a 138±73.2 103±22.5
5.6 Discussion This study was the first to assess the capacity of micron-sized thiol functional cross-linked acrylamide polymer hydrogel particles (X3) to bind metals in contaminated waste rock and tailings, and to increase water holding capacities of the substrates. It demonstrated that addition of novel X3 led to improved seed germination in both phytotoxic mine substrates.
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In this study, plant available concentrations of Al and Cu in mine waste rock were significantly reduced by addition of X3 to below the phytotoxicity thresholds for both native metallophyte grasses A. lappacea and A. scabra, thereby allowing seeds to germinate. This has previously been shown to be due to the irreversible adsorption of these metals by the thiol functional group of X3 (Rossato et al., 2011). The increase of pH in the waste rock amended with X3 might also have contributed to the reduction of metal availability to seeds, as Al and Cu are more available in low pH conditions (Adriano, 2001, Kikui et al., 2005, Meda and Furlani, 2005). Falatah et al. (1996) also reported increased pH of soil with the addition of 0.2 to 0.6% (w/w) of polymers (Broadleaf P4, Agrihobe, Aquasorb and hydrogel). Other studies have investigated the use of micron-size polymers for mine soil remediation. For example, Qu and de Varennes (2010) reported a considerable decrease of water-extractable Pb, Cu and Zn in mine soils following application of cross-linked polyacrylate hydrogel polymer. However, in Pb contaminated soil, the metal-binding capacity of the polymer was found to compete with its water holding capacity which decreased progressively as the polymer sorbed Pb (Guiwei et al., 2008). Activities of soil enzymes such as urease were also impaired by the addition of the polymer, which suggested alteration of soil microbial functionality (Guiwei et al., 2008). Addition of polymer also promoted the formation of microsites rich in water and nutrient, which were favourable to roots and microorganisms (Qu and de Varennes, 2010).
X3 remarkably increased water holding capacities of both waste rock and tailings. This increase is crucial as it could provide water to plants in drought conditions by slowly releasing water to the roots (Rossato et al., 2009) and enhancing plant establishment, particularly in metal contaminated sites (Mendez and Maier, 2008a). Previous studies demonstrated that the water absorbing capacity and the degree of swelling of hydrogels decreased with increasing salinity (Hussien et al., 2012, Qu and de Varennes, 2010). Our study shows that the effectiveness of X3 particles in increasing WHCs of the extremely saline waste rock and tailings was not affected by the presence of salt. Abd El- Rehim et al. (2006) studied the absorbency of superabsorbent hydrogel (polyacrylamide/potassium polyacrylate) in different salt solutions and showed that the absorbent capacity of the hydrogel decreased rapidly at lower pH (4 - 2). In the current study, pH did not affect the capacity of X3 to increase the water holding capacities of waste rock and tailings at or above pH 3.2.
Our study shows that, with the addition of X3 particles, seeds of A. lappacea and A. scabra were able to germinate on extremely saline and metal contaminated waste rock and tailings (which were previously phytotoxic). Seed germination and root elongation have been used in the past to evaluate toxicity of nanoparticles to higher plants (Ma et al., 2011) as this is the first physiological stage of
174 plant growth affected by metals (Shanker et al., 2005). In contrast with other nanoparticles (which had toxic effects on seed germination, such as zero-valent iron and nanosilvers on shoot growth of ryegrass, barley and flax (El-Temsah and Joner, 2012), nano-scale Fe3O4, TiO2 and carbon particles on root growth of cucumber (Mushtaq, 2011) and nanomaterial quantum dots on germination of rice seeds (Nair et al., 2011)), our study confirms that the X3 particles are not toxic to seed germination and supports the previous findings by Rossato et al. (2011). Addition of X3 to waste rock and tailings significantly increased germination percentages, root length and RTI of both grass species which suggests alleviation of toxicity in amended substrates (Torres and de Varennes, 1998). Enhanced germination of seeds and radicle elongation in the presence of X3 hydrogel polymer particles under water restricted conditions confirm the ability of X3 to store and progressively release water to plant roots and improve plant water relations under drought conditions. This is consistent with other studies on hydrogels which have been reported to stimulate root proliferation and aggregation, leading to increased contact with moisture and improved water use efficiency (Agaba et al., 2011) which in turn increased shoot and root biomass (Davies and Castro-Jimenez, 1989, Dorraji et al., 2010). Addition of acrylamide based hydrogels to sandy soils for cultivation purposes was also found to improve water retention of the soil matrix, reduce plant watering frequency and increase plant growth and performance (Abd El-Rehim et al., 2004). Additional benefits of amending soils with acrylamide based polymers include increased (Blodgett et al., 1993) and prolonged (Huttermann et al., 1999) plant water availability when irrigation ceased, reduction of soil erosion, water run-off and compaction, and increased aeration, microbial activity and plant survival (Abd El-Rehim et al., 2004, Guiwei et al., 2008).
5.7 Conclusion Our study shows that X3 micron-size particles significantly ameliorated the quality of mine waste substrates so as to alleviate metal toxicity to radicles and provide moisture in water restricted conditions. This led to improved imbibition and the subsequent germination and radicle elongation of A. lappacea and A. scabra seeds in otherwise phytotoxic mine waste rock and tailings. Further research will investigate interactions between the X3 polymer, water and ions in a long-term glasshouse experiment on contaminated mine wastes and their effects on plant establishment in phytotoxic conditions. Sustainability of the X3 particles in mine spoils will also be tested.
Acknowledgements The study was funded by a Pathfinder grant from UniQuest Pty Ltd (The University of Queensland, Australia). The authors are grateful to Mr Craig Claremont, summer scholarship student recipient at
175 the University of Queensland, for his assistance with the experimental work. The assistance from mine site personnel and comments of the anonymous reviewers are gratefully acknowledged.
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Chapter 6. General Discussion and Conclusions
Commercial mineral extraction from the Earth’s crust (mining) occurs where the concentrations of desired materials (especially metals, some non-metals, and carbon) are much higher than in most rocks and soils that cover the Earth’s surface. These mineralised areas usually undergo extensive disturbance during the mining process, and the extraction of the desired materials is rarely complete, so that the post-mining landscape usually contains novel landforms and elevated concentrations of these materials or of their breakdown products, which potentially affect the immediate soil environment but may also affect more distant locations through transport in the water, redistributed sediments or in the air.
Common requirements for the relinquishment of mining leases are the avoidance of discharge of defined materials from the site and the establishment of defined ecosystems or vegetation assemblages on the reconstructed landforms. The process of vegetation establishment depends on the existence of suitable conditions in the substrate that permit plant growth including its physical structure and texture, plant water availability, pH, the absolute and relative concentrations of nutrient elements and the concentrations of non-essential and potentially toxic substances. These requirements mean that the range of soil conditions suitable for plant growth are narrow compared with the range of conditions commonly found in post-mining landscapes. Therefore, plant establishment on these sites depends on modification of soil conditions so that desired species can be maintained, or on the selection of species that are relatively tolerant of the new conditions. Both lines of activity are necessary for the attainment of acceptable ecosystem stability at an acceptable cost to the mining leaseholders and subsequent land managers.
This thesis addresses both of these lines of activity. The general objective is to: (1) examine the relationship between selected Australian plant species and the properties of metal-rich or contaminated sites, and (2) test the contributions of synthetic designer metal-binding particles and these species to site remediation. The two components of this objective were investigated by considering:
(1) Plant properties under metal conditions: The aims were to:
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(a) develop a new conceptual framework to characterize plant responses to soil metals based on metal transporter kinetic parameters, and validate this model using field survey of Queensland native plant species and literature data. (b) examine the effect of arsenic (As(V)), copper (Cu), zinc (Zn), manganese (Mn) and lead (Pb) on the germination characteristics of potential native metallophytes, including Astrebla lappacea, Austrostipa scabra, Themeda australis and Acacia harpophylla in the absence of metal-binding particles.
(2) Modification of conditions to enhance plant establishment on contaminated substrates: The aims were to: (a) assess the potential of metal-binding particles to sequester metals in contaminated substrates and to increase substrate water holding capacity. (b) investigate the effect of metal-binding particles on germination of selected native metallophyte grasses (A. lappacea and A. scabra as model species) on contaminated substrates. (c) examine the effect of metal-binding particles on plant establishment and growth of the native metallophyte grass A. lappacea on contaminated substrates under controlled conditions in a short-term glasshouse trial.
6.1 Objective 1. Examine the relationship between selected Australian plant species and the properties of metal-rich or contaminated sites 6.1.1 To develop a new conceptual framework to characterize plant responses to soil metals based on metal transporter kinetic parameters, and validate this model using field survey of Queensland native plant species and literature data. Increasing mining activities in recent decades have resulted in large areas of contaminated land which require rehabilitation. Phytoremediation is a cost effective and environmentally benign biological remediation technology which involves plants for rehabilitation of environmentally sensitive contaminated sites (Hartley and Lepp, 2008). Native plants, including Australian native plant species, have been widely used for phytoremediation, through either phytoextraction or phytostabilization (Barrutia et al., 2011, Tongway and Ludwig, 2011, Testiati et al., 2013), but little is known about their responses and mechanisms of tolerances to metals.
In Chapter 3, a new conceptual framework was developed for characterization of plant responses to soil metals based on metal transporter kinetic parameters. The model was developed from extensive literature studies that sought to understand mechanisms of plant response to metals, such as how metals are taken up from soils by roots and transferred into upper plant parts and how metals are distributed within plant parts. Published data showed that the patterns of uptake and responses to 178 metals in plants varied with metal concentrations in substrates, but were, in principle, similar for all examined plant species. At low available external metal concentrations, uptake rates of metals display hyperbolic curves with respect to increasing metal concentration (Michaelis-Menten kinetics) and uptake is facilitated by high affinity metal transport systems (HATS), whereas at high external metal concentrations, the uptake rate is linearly related to metal concentration and uptake is mediated by low affinity metal transport systems (LATS) (Cohen et al., 2004, Malagoli et al., 2004, Pedas et al., 2005, Ghaly et al., 2008, Bravin et al., 2010). Field survey data of Queensland native plant species collected from metal-enriched areas and the literature data confirmed and validated the metal uptake pattern presented in Chapter 3. The finding of common uptake processes differs from the widely accepted classification of metallophyte plant responses to metals established by Baker (1981) who proposed that accumulators and hyperaccumulators display a hyperbolic relationship between plant shoot and soil metal concentrations, and indicators a linear response, while in excluders, plant metal concentration remains constant and low over a wide range of soil metal concentrations. The new framework of plant responses to metals developed in this thesis presents a comprehensive and unified classification of plant-soil metal relationships as it considers mechanisms involved in metal transport and distribution within plants. It also considers the importance of plant available metal concentration in substrates.
The framework established in Chapter 3 is based on the proposition that relevant metal concentrations in soils are the forms of metals that are readily available for plant uptake. Therefore, responses of plants to metals in soils are best related to available forms of metals in soils instead of total metal concentrations in soils. Field surveys on Queensland native plant species showed clear relationships between aluminium (Al), iron (Fe), manganese (Mn), nickel (Ni) and copper (Cu) concentrations in plant parts and the corresponding plant available metal concentrations in soils, whereas there was no relationship with total soil metal concentration. The findings are in agreement with previous reports, that plants absorb heavy metals that are present as soluble components in the soil solution and readily available for plant roots (Chibuike and Obiora, 2014, Garcia-Gomez et al., 2014, Saha et al., 2017). Although some studies of relationships between plant and soil metal concentrations and for the identification of metallophyte plants (e.g. hyperaccumulators) have focused on plant available forms of metals in soils, the total concentrations of metals in soils are still being widely used. This could be due to the legacy of Baker’s (1981) classification of plant responses to metals, which was based on total metal concentrations in soils, and is still the classification most widely used for the identification of metallophyte plants (especially hyperaccumulators). The new framework developed in Chapter 3 provides a more satisfactory
179 explanation of the relationships between metals in plants and in soils by considering the processes involved in plant metal uptake.
Based on the new framework in Chapter 3, new terminology for metallophytes was proposed. Metallophytes are identified as metal tolerators; either shoot/leaf tolerator, root tolerator or both, depending on distribution of metals within plant parts. Kinetic parameters proposed in Chapter 3 enable identification of tolerators based on whether the tolerator is grown on low (maximum uptake, Vmax; and metal affinity, Km of HATS) or high plant available metal concentration (LATS slope). Vmax and Km have been mostly used to determine metal uptake by plants in solutions. However, field data of Queensland native species (Chapter 3) and some literature data (Kozhevnikova et al., 2014) demonstrated that the parameters were also applicable for determination of plant response to metals in soils. In contrast, LATS slope has received little attention in the literature. Nevertheless, the study in Chapter 3 showed that LATS slope was a good and sufficient sole parameter for the determination of plant responses in metal-enriched soils while the intercept of the uptake equation did not show a close correlation between metals in plants and in soils. The framework also considers the transition metal concentration between the high- and low- affinity transport systems, which allows us to clearly define which parameters are to be used for the classification of metal tolerators. This more detailed information on a species’ type of response to a particular metal and concentration range could deepen the understanding of species’ adaptability to different soil conditions.
Thus, the framework in Chapter 3 could assist, through more comprehensive identification of responses of metallophyte species to different soil metal concentrations, determination of the suitability of a species for different phytoremediation objectives. For a more complete identification using the framework, a complete set of field data is required to be able to adequately classify metallophyte plants into shoot/leaf tolerator or root tolerator or shoot/leaf and root tolerator. This means that plant parts (shoot/leaves and roots) and corresponding soil samples are crucial for studying the plant-soil-metal relationship. As the framework could also apply to any elements or plant types or plant parts, it could be beneficial to a wider range of interest groups.
6.1.2 To examine the effect of arsenic (AsV), copper (Cu), zinc (Zn), manganese (Mn) and lead (Pb) on the germination characteristics of potential native metallophytes including Astrebla lappacea, Austrostipa scabra, Themeda australis and Acacia harpophylla in the absence of metal-binding particles. Despite limited information on metal tolerance characteristics of Australian native plant species, they have been used widely for rehabilitation of metal contaminated sites. In order to effectively 180 utilize any species for phytoremediation, an understanding is required of germination and plant growth responses, toxicity thresholds and tolerance mechanisms to metal/loids, and the soil or substrate characteristics affecting plant responses to metals (Reichman, 2002, Krämer, 2010).
Germination is the first and critical stage of plant life and is considered more sensitive to metal toxicity than later stages (He et al., 2014, Madejón et al., 2015). In Chapter 4, firstly, several plant metal/loid tolerance indicators were evaluated for the germination of selected Australian native species. A detailed study was important because different indicators may show different levels of sensitivity to metal/loid toxicity, which would lead to different reported plant metal tolerance thresholds. The results of Chapter 4 demonstrate that the order of indicator sensitivity to metal/loids for three grass and one tree species was: radicle tolerance index (RTI) > maximum germination percentage (Gmax) > shoot tolerance index (STI) > mean germination time (MGT). Higher proportions of germinated A. lappacea seeds lacked a radicle in As and Pb treatments (Chapter 4), contributing to the relative sensitivity of RTI in response to metal/loids toxicity. Previous studies on various metal’s toxicities during germination have reached similar conclusions (He et al., 2008, Almas et al., 2012, Hou et al., 2014). It appears that, regardless of metal identity or plant species, roots are the first of all plant parts to be affected by toxic elements in media/soils. Radicles/roots are important organs for uptake of nutrients and water for seed development and growth (Xiong, 1998, Hou et al., 2014). Therefore, plant growth and establishment would fail if roots do not develop. The higher sensitivity of radicles/roots to toxic concentrations of metal/loids, especially to non-essential elements, and the importance of roots for plant growth, suggest that radicle/root tolerance index could be considered as the key indicator for establishment of metal toxicity thresholds in plant species.
In Chapter 4, tolerance thresholds for germination of A. lappacea, T. australis, A. scabra and A. harpophylla to different metal/loids were also defined. Based on radicle tolerance index (RTI) as the key tolerance indicator, the species showed different degrees of tolerance to metal/loids during germination, but they could all be classified as metallophytes, confirming their current suitability and use for mine site rehabilitation (Huxtable, 2003, Huxtable et al., 2005). Although the thresholds established in this study were only for the germination stage, they may provide key information for studying species’ metal tolerance characteristics in later stages of plant growth, although further experimental work would be required to ascertain that toxicity thresholds established during seed germination are also applicable to plant growth.
Arsenic is not an essential element and is toxic to plant growth even at low concentration (Bhupendra et al., 2014). The lowest concentration of As tested in Chapter 4 (13.34 μM) was 181 considered as a critical concentration to a wide range of European species (Porter and Peterson, 1977). Acacia harpophylla and A. scabra were found to be extremely tolerant to high concentration of As, with tolerance thresholds >667.36 μM. Astrebla lappacea were tolerant to high levels of As with a tolerance threshold between 400.41-533.88 μM. In contrast, T. australis was the least tolerant to As, with a tolerance threshold of 66.73-133.47 μM, although it has been widely used for rehabilitation of mine sites in Australia. Tolerance levels of A. lappacea and A. scabra were higher than those of previously studied species such as Phaseolus aureus, Helianthus annuus L., and Hordeum vulgare L. (Kaur et al., 2012, Imran et al., 2013, Sanal et al., 2014).
Copper, Zn and Mn are essential elements and are important for plant growth at low concentrations. However, the lowest concentrations of these metals tested in Chapter 4 (0.5, 9, 8 µM, respectively) were considered critical for Australian native trees during seedling growth (Reichman, 2001) but A. lappacea and T. australis were both tolerant to Cu and Zn (RTI toxicity threshold >200 µM and 500 µM, respectively), while both showed a similar tolerance level as A. harpophylla to Mn (RTI toxicity threshold >10240 µM). Astrebla lappacea was able to germinate and establish on highly phytotoxic waste rock and tailings in the presence of metal binding particles (Chapter 5 and Appendix IX).
Lead is also toxic to plants even at lower concentration. Austrostipa scabra was superior to A. lappacea and T. australis in its ability to tolerate high concentrations of Pb, with a tolerance threshold >9600 μM (Chapter 4). Astrebla lappacea Pb tolerance threshold was between 3840-4800 μM, whereas T. australis had a tolerance threshold <240 μM. These findings indicate that A. scabra and A. lappacea were tolerant to different concentrations of Pb during germination. A previous study on toxicity of Pb to A. lappacea by Rossato et al. (2011) demonstrated that A. lappacea was unable to germinate on 9600 μM Pb unless metal-binding particles (X3) were present.
Toxicity thresholds of the native species established in Chapter 4 could contribute to the understanding of responses of Australian native plant species to metal/loids. Their ability to tolerate drought and harsh environments of Australia has been reported by various researchers (Orr and Evenson, 1993, Windsor and Clements, 2001, Huxtable et al., 2005, Munksgaard and Lottermoser, 2013). However, their metal tolerance thresholds had never been studied. Therefore, the thresholds established in Chapter 4 could contribute crucial information on tolerance characteristics of the species during germination and their suitabilities for mine land rehabilitation, although further research would be required to confirm that toxicity thresholds established during species germination also apply to plant growth. Utilization of native plant species for mine land rehabilitation could minimize the risk that potentially invasive introduced species may compromise 182 rehabilitated metal contaminated sites in Australia. In addition endangered native species, such as A. harpophylla, may be conserved, thereby enhancing regional plant biodiversity (Huxtable et al., 2005).
6.2 Objective 2. Modification of conditions to enhance plant establishment on contaminated substrates 6.2.1 To assess the potential of metal-binding particles to sequester metals in contaminated substrates and to increase substrate water holding capacity. Mining is considered to be a major cause of metal contamination in soils and is the major source of transportable contaminants (Selinus, 2013, Xiao et al., 2017). Mine waste materials normally contain high concentrations of metals with other various factors limiting plant growth. Characterization of mine waste rock and tailings in Chapter 5 and Appendix IX demonstrated that both materials were not favourable for plant growth, even for metallophyte plants, because the materials contained phytotoxic concentrations of metals, were highly saline and acidic (for waste rock only). These limiting conditions are often identified in mine sites and are considered one of the major problems for rehabilitation of metal contaminated sites (Alvarez et al., 2003, Mendez and Maier, 2008a).
Nanotechnology could offer an alternative and cost effective solution for decontamination of metal contaminated sites (Karn et al., 2009). It has been used for various purposes, including site treatment and remediation (Bidar et al., 2007, Alqudami et al., 2012, Khin et al., 2012, Klimkova et al., 2011, Shah et al., 2016) and could complement existing techniques (Sun et al., 2006).
In Chapter 5, specifically designed micron-sized thiol functional cross-linked acrylamide metal- binding polymer particles (X3) greatly improved the biological quality of mine waste rock and tailings by reducing concentrations of toxic elements below their phytotoxic thresholds and allowing the native grass species A. lappacea and A. scabra to germinate. The result supports a previous study by Rossato et al. (2011), who reported significant decreases in initial soluble concentrations of Zn2+, Cu2+ and Pb2+ in the presence of X3 via strong selective interactions between the thiol binding groups and the heavy metals. The effectiveness of xanthate nanoparticles in selectively removing heavy metals (alkali metals have only a low affinity for binding) from water to concentrations below 1 ppm has also been reported (Chaudhari and Tare, 1999, Bell et al., 2006, Chang et al., 2007, Xu et al., 2012). The capacity of metal-binding particles (X3) to bind different metals at the same time in mine waste materials suggests that the particles were not metal specific. Therefore, X3 could be a potential solution for remediation of mine land or metal contaminated
183 sites which are often polluted by not only one toxic element but various hazardous materials (Mendez and Maier, 2008a).
Chapters 5 also demonstrated that X3 was able to greatly increase water holding capacity of waste rock and tailings. The finding is also in agreement with previous studies (Rossato et al., 2011, Jatav et al., 2013, Zhou and Chen, 2017). Although waste rock and tailings were extremely saline, the effectiveness of X3 for increasing water holding capacity was not affected by salinity. Thereby, it allowed germination of A. lappacea and A. scabra on the highly phytotoxic saline substrates. With this quality, X3 particles proved to be potentially useful at mine sites which are often characterized by lack of water (drought). Although metal-tolerant plants have been used to rehabilitate mine sites, limited soil water contents at many Australian locations often prevent seed germination. Seedling emergence requires water for growth and plant development and most seeds germinate on topsoil with a relatively high plant available water content (Muñoz et al., 2016). In water limiting conditions, slow release of water by X3 particles could potentially assist seeds to successfully germinate and seedlings to become established.
With the capacity to both bind metals and supply water, X3 particles were proven to be potentially beneficial for site remediation. Therefore, X3 could potentially reduce costs of remediation of large- scale contaminated sites, reduce cleanup time, and minimize treatment, disposal quantities and costs for contaminated sites (Karn et al., 2009). The possible increase in incidence of drought due to future climate change may adversely affect seedling establishment and biodiversity of native plant communities (Karn et al., 2009). The capacity of X3 to absorb water and release it gradually to seeds during germination could become an alternative solution for mine site rehabilitation.
6.2.2 To investigate the effect of metal-binding particles on germination of selected native metallophyte grasses (A. lappacea and A. scabra as model species) on contaminated substrates. Chapter 5 demonstrated that metal binding particles were not toxic to seeds of A. lappacea and A. scabra germinating on highly toxic waste rock and tailings. Overall, maximum germination percentage (Gmax %), mean germination time (MGT) and radicle tolerance index (RTI) of A. lappacea and A. scabra in X3-amended substrates were higher or not significantly different from unamended substrates. In addition, radicles of germinated seeds of A. lappacea and A. scabra in X3-amended substrates did not show any toxicity symptoms (Chapter 5). Rossato et al. (2011) also reported that X3 particles had no negative physiological effects on seed germination of A. lappacea. Stampoulis et al. (2009) showed carbon nanotubes (MWCNTs) had no adverse effect on germination and root length of zucchini species. In contrast, other nanoparticles such as Cu- 184 nanoparticles greatly reduced radicle elongation of zucchini and maize (Stampoulis et al., 2009, Wang et al., 2012, Yang et al., 2017). Roots have been used as an indicator of nanoparticle toxicity in higher plants (Rossato et al., 2011). Thus, absence of toxicity symptoms in radicles of germinating seeds indicates that X3 particles were harmless to seed germination of A. lappacea and A. scabra, and its presence promoted germination of seeds and radicle elongation of these native species on highly phytotoxic mine waste materials.
Seed-based plant restoration is one option for restoration of degraded ecosystems, at both small and large scales (Broadhurst et al., 2016). Although a large-scale sowing of native seeds is a preferred remediation technique, it is often less successful (LPSDP, 2006). The X3 particles could be potentially beneficial by providing a more benign planting condition and therefore increase the likelihood of seed germination and seedling emergence and establishment. In addition, X3 particles could be an additional source of moisture for germination which often fails due to lack of water in drought conditions (Muñoz-Rojas et al., 2016).
Potential use of X3 particles in mine sites could also be beneficial for the conservation of native plant species. The characteristics of the X3 particles, which were not toxic to seed germination of Australian native grass species, and its potential use in drought-susceptible areas and mine sites, could promote biodiversity of native species and therefore minimize the use of introduced species which may be harmful for local ecosystems. In addition, the use of X3 particles for improving soil condition could also assist in conserving threatened native plant species.
6.2.3 To examine the effect of metal-binding particles on plant establishment and growth of the native metallophyte grass A. lappacea on contaminated substrates under controlled conditions in a short-term glasshouse trial. X3 greatly improved chemical and physical qualities of phytotoxic waste rock and tailings resulting in improved seed germination and healthy early establishment of A. lappacea (Appendix IX).
Mine waste rock is discarded in the process of accessing the valuable ore, and tailings, which are the residue after extraction of minerals from the ore, are significantly finer grained (silt or sand size) than waste rock and are considered to be the major source of pollution at mine sites (Mendez and Maier, 2008b). One of the challenges for mine site rehabilitation is the presence of highly toxic materials in mine lands. Mine waste rock and tailings used in Appendix IX showed similar detrimental characteristics to plant growth, as has previously been reported. At Mt. Morgan and Mt. Isa mine sites (Queensland, Australia), various attempts at site rehabilitation (through plant establishment) have not been successful due to the presence of highly toxic materials and physical
185 characteristics that limit plant growth, such as soil surface compaction and lack of moisture (Appendix IX).
Studies in Chapter 5 and Appendix IX demonstrated that adding metal-binding particles (X3) to mine waste rock and tailings greatly reduced availabilities of aluminium (Al), copper (Cu), zinc (Zn), manganese (Mn) and cobalt (Co) in waste rock, and Zn, Mn, Co and cadmium (Cd) in tailings, from initially phytotoxic concentrations to below phytotoxic thresholds, allowing healthy plant growth (Mulvey and Elliott, 2000). The findings supported a previous study by Rossato et al. (2011) who observed that X3 significantly reduced concentrations of Pb, Cu and Zn in solutions. Although Cu, Zn, Mn and Co are considered essential nutrients for plant growth at low concentrations, high concentrations of these metals are detrimental to plants, while Al and Cd are not essential elements and are toxic to plant growth at low concentrations. Previous studies have observed various toxicity effects of these elements on different stages of plant growth (Dukic et al., 2014, Adrees et al., 2015, Imadi et al., 2015). Reduction of toxic element concentrations in waste rock and tailings in the presence of X3 particles allowed A. lappacea to germinate, emerge and establish on the otherwise phytotoxic waste materials (Appendix IX). Astrebla lappacea is a native Australian metallophyte species and has been used for rehabilitation of metal contaminated sites (Huxtable, 2000). The species was also tolerant to normally toxic concentrations of As(V), Cu, Zn, Mn and Pb in solutions during germination (Chapter 4). Nevertheless, without prior amelioration of waste rock and tailings with X3 particles, this metallophyte species was not able to establish on highly toxic substrates (Appendix IX).
X3 particles also greatly increased water holding capacity and plant available water (PAW) of waste rock and tailings (Appendix IX). The capacity of X3 particles to increase water holding capacity had been previously shown by Rossato et al. (2011). In mine spoils where shallow intact bedrock and large-sized waste material particles are present, water availability to plants is often insufficient to sustain plant growth during drought (Sheoran, 2010). X3 could potentially ameliorate such condition as it could release water slowly to plants. In extremely saline materials such as waste rock (EC 2.87±0.03 dS.m-1) and tailings (EC 5.54±0.02 dS.m-1), osmotic potential can have a great effect on total water potential and leads to low water availability (The State of Queensland, 2011). Addition of X3 alleviated salinity and improved water availability for the establishment of A. lappacea. The study is in agreement with Hojjat and Kamyab (2017) who reported silver nanoparticles improved water availability to germinating Fenugreek seeds under salinity stress. In contrast, the capacity of some gel-forming polymers to absorb water may be hampered by salinity (Johnson, 1984). In metal contaminated mine lands, where salinity is one of the limiting factors to
186 plant establishment, X3 particles could potentially be used to minimize the negative effect of salts on plant growth, especially during germination and early plant establishment, and provide water by slowly releasing it to plants. For rehabilitation of coal mine spoils in Australia, where salinity is the greatest concern (Li et al., 2014), X3 particles could be an alternative substrate amendment.
Mine soils are commonly highly compacted, characterized by bulk density >1.7 g/cc, which limits plant growth because most species are unable to extend roots effectively through these soils (Sheoran, 2010). This phenomenon was also observed in waste rock and tailings used for the experiment described in Appendix IX. Besides containing highly toxic materials, the surface hardness of waste rock and tailings prevented radicles of germinated seeds from penetrating into unamended substrates and seedlings were unable to emerge and establish. Addition of X3 particles greatly reduced surface compaction, thus enhancing root penetration of A. lappacea during early establishment on both substrates.
In addition, X3 was not toxic to A. lappacea plants, as no toxicity symptoms were observed in leaves of plants growing on substrates amended with X3, as compared to the control. This means the particles are safe for plants, as they have no phytotoxic properties, and have valuable remediation properties.
The X3 technology cost will depend on the level of contamination at sites and method of application employed. The cost should be in the range of other polyacrylamide-based soil conditioners, which have recently been applied widely for degraded land management because they were shown to be sustainable, environmentally friendly, highly efficient and inexpensive (Orts et al., 1999, Sojka et al., 2007, Guiwei et al., 2008, Hüttermann et al., 2009), at costs estimated at US$ 2-4 kg-1 (Hüttermann et al., 2009) or € 2 kg-1 (Guiwei et al., 2008).
In short, X3 particles could be a safe and cost-effective solution for rehabilitation of mine lands as they could ameliorate chemically and physically unfavourable mine waste materials.
6.3 Concluding remarks and future research With requirements for more environmentally friendly and sustainable ways of rehabilitating metal contaminated sites, demands for application of phytoremediation using native plants are increasing. However, availability of plants for different types of phytoremediation is still limited due to lack of information on plant metal response characteristics, including their metal tolerance, and contaminated soils often require amelioration prior to revegetation.
187
The results of this thesis provide a comprehensive framework for characterization of plant responses to soil metals which considers relevant aspects in the plant-soil-metal relationship, including metal transporter kinetic parameters and plant available concentrations of metals in soils. Therefore, the framework could directly assist relevant interest groups in application of a more concrete phytoremediation technique to metal contaminated areas through selection of appropriate metal tolerators for specific site problems or phytoremediation objectives.
The thesis also establishes metal/loid tolerance thresholds during germination of some Australian native plant species which have been frequently used in mine sites. This information could contribute to a more effective use of the species for mine site rehabilitation and could serve as a foundation for establishment of metal/loid tolerance thresholds of the species at the next growth stage, i.e. during plant establishment. The methodology followed could also be used as a basis for future research investigating plant tolerance to metals during germination of other Australian native species, for which there is currently very limited knowledge, to reduce this gap of knowledge and help select suitable species for phytoremediation. Use of suitable native plant species to remediate metal contaminated sites would assist in the conservation of endangered plant species and therefore promote and protect plant biodiversity and site ecosystems.
The thesis also proposed a new remediation technology (X3 particles) that is able to ameliorate metal contaminated sites prior to revegetation. X3 has the potential to become a cost-effective and safe replacement to current expensive, ineffective and environmentally disruptive remediation technologies available on the market, such as excavation and disposal to landfill. X3 particles could contribute to a more sustainable way of managing and rehabilitating metal contaminated sites in combination with native metal tolerator species, which would encourage and promote the application of phytoremediation techniques in site remediation.
This research is a significant and novel contribution to a wide interest group, particularly in the area of soil remediation, where available information and knowledge on plant metal response (plants in general and Australian native species in particular) are still limited. In addition, this research is the first proof of concept study establishing the short-term effects of X3 on improvement of chemical and physical properties of contaminated substrates, and ameliorated seed germination and healthy early plant establishment on otherwise phytotoxic substrates. Although the X3 technology seems very promising, further studies are needed in order to examine the long-term effects of the X3 particles on waste materials and plant establishment in glasshouse and field conditions. This would enable identification of the most effective methods of X3 application in field conditions to suit the
188 needs of specific rehabilitation purposes. In addition, further studies on the effects of the X3 particles on soil microbes is also needed in order to ensure that X3 would not be toxic to soil biota.
189
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Appendix I. Location details of plant and soil sample collections. Annual Number Altitude Location Soil type/Site description Site Latitude Longitude rainfall of species (m) (mm) collected Ridgelands, 600 m from Atkinson Rd Turnoff from Bruce Highway, 1 S23˚05’35” E150˚16’20” 11.3 946.7* 9 Rockhampton 40 km NW of Rockhampton, 50 km SE of Marlborough. and Flat country on Serpentinite; Eucalyptus fibrosa woodland Marlborough 44km SE of Marlborough, 46km NW of Rockhampton. 2 S23˚01’22” E150˚16’38” 11.3 946.7* 5 Base of a rocky hill; skeletal soil; Eucalyptus fibrosa/ Corymbia xanthope woodland with understorey of Acacia leptostachya and other shrubs, ground stratum dominated by Triodia mitchellii 48 km SE of Marlborough, 45 km NW of Rockhampton. 3 S23˚01’34” E150˚17’54” 11.3 946.7* 1 Disturbed area on roadside; landform - flat Roadside - Old Byfield Road, 27 km NE of Rockhampton 8 S23˚09’25” E150˚38’05” 11.3 946.7* 9 and 12 km WSW of Yeppoon, 2.3 km SE of MT Cobbera. Mid slope of low hill; Eucalyptus fibrosa grassy woodland with understorey dominated by Acacia 5 km SW of Marlborough, 3.3 km E of Mt Slopeaway. Flat 4 S22˚51’06” E149˚52’07” 11.3 946.7* 5 area on alluvium derived from Serpentinite. Old Bruce Highway, 8.5 km SW of Marlborough, 2 km S 5 S22˚52’15” E149˚50’12” 11.3 946.7* 7 of Mt Slopeaway. Hills on Serpentinite; Eucalyptus fibrosa woodland with dry rainforest species in the understorey 11 km S of Marlborough, 3.7 km S of Mt Slopeaway. Flat 6 S22˚52’16” E149˚48’24” 11.3 946.7* 1 area on alluvium derived from Serpentinite; grassy cleared area with Melaleuca bracteata and Melaleuca linariifolia, previously dominated by Eucalyptus tereticornis Old Bruce highway, adjacent to Marlborough Creek, 4.6 7 S22˚50’57” E149˚42’02” 11.3 946.7* 1 km SW of Marlborough, 3.3 km E of Mt Slopeaway. Flat area on alluvium derived from Serpentinite; cleared area with a few remnant trees – Eucalyptus tereticornis/ Corymbia dallachyana 250
Mt. Morgan tailings dam Mor S23o38’28” E150 o22’31” 240 740* 2 mine site Cracow mine I/Q dump at GP South/Eastern side WTP S25o17’0” E150o18’0” 317 677.5* 3 site
*Source: http://www.bom.gov.au/climate/averages/tables/cw_039082.shtml http://en.wikipedia.org/wiki/Rockhampton http://en.wikipedia.org/wiki/Marlborough,_Queensland http://www.bonzle.com/c/a?a=p&p =256676&wetgr=r&d=w&c=1&x=150.29426&y=-25.2835&w=20000&mpsec=0
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Appendix II. Maximum germination percentage (Gmax) (shoot only or with a radicle) of A. lappacea, T. australis, A. scabra and A. harpophylla exposed to various concentrations of As(V) solutions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean for five (A. lappacea and A. harpophylla) or eight (T. australis and A. scabra) replicates. st Asterisks indicate when Gmax was significantly different (P<0.05) from the control. ‘1 exp.’ refers to ‘first experiment’; ‘2nd exp.’ refers to ‘second experiment’.
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Appendix III. Maximum germination percentage (Gmax) (shoot only or with a radicle) of A. lappacea, T. australis and A. harpophylla exposed to various concentrations of Cu solutions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean for five
(A. lappacea and A. harpophylla) or ten (T. australis) replicates. Asterisks indicate when Gmax was significantly different (P<0.05) from the control. ‘1st exp.’ refers to ‘first experiment’.
253
Appendix IV. Maximum germination percentage (Gmax) (shoot only or with a radicle) of A. lappacea, T. australis and A. harpophylla exposed to various concentrations of Zn solutions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean for five
(A. lappacea and A. harpophylla) or ten (T. australis) replicates. Asterisks indicate when Gmax was significantly different (P<0.05) from the control. ‘1st exp.’ refers to ‘first experiment’.
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Appendix V. Maximum germination percentage (Gmax) (shoot only or with a radicle) of A. lappacea, T. australis and A. harpophylla exposed to various concentrations of Mn solutions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean for five
(A. lappacea and A. harpophylla) or ten (T. australis) replicates. Asterisks indicate when Gmax was significantly different (P<0.05) from the control. ‘1st exp.’ refers to ‘first experiment’; ‘2nd exp.’ refers to ‘second experiment’.
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Appendix VI. Maximum germination percentage (Gmax) (shoot only or with a radicle) of A. lappacea, T. australis and A. scabra exposed to various concentrations of Pb solutions. Vertical bars when larger than the symbol indicate ± standard error (SE) of the mean for five (A. lappacea) or eight (T. australis and A. scabra) replicates. Asterisks indicate when Gmax was significantly different (P<0.05) from the control.
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Appendix VII. Effect of various concentrations of As(V), Cu, Zn, Mn and Pb solutions on mean germination time (MGT) of Astrebla lappacea, Themeda australis, Austrostipa scabra and Acacia harpophylla. MGT (hours or days)* Metal/loid A. lappacea T. australis A. scabra A. harpophylla MGT R MGT S MGT R MGT S MGT R MGT S MGT R MGT S
days days days days days days hours hours 1st exp.
Control (SDIW) 4.5±0.6a 4.2±2.2a 5±0.9a 0±0a 3.8±0.2a 0±0a 33.1±2.3a 0±0a As(V) (13.34 μM) 5.7±1.1a 3.2±0.7a 3.7±0.7a 0±0a 3.7±0.3a 0±0a 37.3±2.9a 0±0a As(V) (66.73 μM) 3.5±0.3a 6.4±0.9a 4±0.5a 0±0a 3.6±0.4a 0±0a 37.1±2.1a 0±0a As(V) (133.47 μM) 3.8±0.1a 6.5±0.9a 2.3±0.2b 0±0a 4±0.3a 0±0a 40.3±3.7a 0±0a As(V) (266.94 μM) 4±0.5a 4.5±0.3a 3.5±0.5a 0±0a 4.7±0.4a 0±0a nt nt As(V) (400.41 μM) 4.3±0.7a 5±0.5a 3±0.3b 0±0a 4.4±0.3a 0±0a nt nt As(V) (533.88 μM) 3.5±0.3a 5.2±0.4a 2.5±0.3b 0±0a 4.4±0.6a 0±0a nt nt As(V) (667.36 μM) 3.3±0.3a 5±0.4a 3.8±0.8a 0±0a 4.1±0.5a 0±0a nt nt 2nd exp.
Control (SDIW) 20.0±1.3a 0±0a
As(V) (133.47 μM) 19.9±2a 0±0a
As(V) (266.94 μM) 21.9±4.3a 0±0a
As(V) (667.36 μM) 20.2±2.4a 0±0a 1st exp.
Control (SDIW) 3.6±0.1a 0±0a 3.9±0.7a 0±0a nt nt 33.8±2.4a 0±0a Cu (0.5 μM) 3.2±0.2a 0±0a 3.1±0.2a 0±0a nt nt 38.6±3.1a 0±0a Cu (10 μM) 3.1±0.2a 0±0a 2.9±0.1a 0±0a nt nt 30.2±1.9a 0±0a Cu (40 μM) 3.9±0.3a 0±0a 3.6±0.3a 0±0a nt nt 33.6±1.7a 0±0a Control (SDIW) 3.1±0.3a 0±0a 3.2±0.2a 0±0a nt nt nt nt Cu (40 μM) 3.2±0.2a 0±0a 2.9±0.2a 0±0a nt nt nt nt Cu (80 μM) 3.2±0.1a 0±0a 2.9±0.2a 0±0a nt nt nt nt Cu (200 μM) 3.5±0.2a 0±0a 3.2±0.2a 0±0a nt nt nt nt
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Appendix VII. contd.
MGT (hours or days)* Metal/loid A. lappacea T. australis A. scabra A. harpophylla MGT R MGT S MGT R MGT S MGT R MGT S MGT R MGT S
days days days days days days hours hours 1st exp.
Control (SDIW) 3.6±0.1a 0±0a 3.9±0.7a 0±0a nt nt 33.8±2.4a 0±0a Zn (9 μM) 2.9±0.1a 0±0a 3.9±0.5a 0±0a nt nt 35±2.2a 0±0a Zn (25 μM) 3.1±0.1a 0±0a 3.3±0.3a 0±0a nt nt 38.6±2.9a 0±0a Zn (100 μM) 3.1±0.2a 0±0a 4.2±0.6a 0±0a nt nt 32.9±2.6a 0±0a Control (SDIW) 3.1±0.3a 0±0a 3.2±0.2a 0±0a nt nt nt nt Zn (100 μM) 3.1±0.2a 0±0a 3±0.1a 0±0a nt nt nt nt Zn (200 μM) 3.2±0.1a 0±0a 3.1±0.3a 0±0a nt nt nt nt Zn (500 μM) 3.2±0.2a 0±0a 2.6±0.2a 0±0a nt nt nt nt 1st exp.
Control (SDIW) 3.56±0.12a 0±0a 3.96±0.79a 0±0a nt nt 33.8±2.4a 0±0a Mn (8 μM) 3.35±0.16a 0±0a 3.1±0.43a 0±0a nt nt 42±3.6a 0±0a Mn (128 μM) 3.47±0.31a 0±0a 4.06±0.66a 0±0a nt nt 37.4±4.1a 0±0a Mn (2048 μM) 3.36±0.26a 0±0a 3.16±0.12a 0±0a nt nt 31.9±2.2a 0±0a 2nd exp.
Control (SDIW) 3.17±0.3a 0±0a 3.26±0.2a 0±0a nt nt 20.04±1.3a 0±0a Mn (2048 μM) 2.97±0.2a 0±0a 2.8±0.3a 0±0a nt nt 18.08±1.4a 0±0a Mn (4096 μM) 3±0.2a 0±0a 2.86±0.2a 0±0a nt nt 15.17±1.4a 0±0a Mn (10240 μM) 2.96±0.1a 0±0a 3.13±0.3a 0±0a nt nt 21.3±2.9a 0±0a Control (SDIW) 4.5±0.6a 4.2±2.2a 3.1±0.4a 0±0a 3.6±0.2a 0±0a nt nt Pb (240 μM) 3.8±0.2a 3.3±0.2a 6.7±0.8b 0±0a 3.1±0.2a 0±0a nt nt Pb (480 μM) 3.7±0.2a 3.5±0.4a 6.5±0.7b 0±0a 3.2±0.2a 0±0a nt nt Pb (960 μM) 4.4±0.4a 4.3±0.9a 7.4±1.3b 0±0a 4±0.3a 0±0a nt nt Pb (1920 μM) 4.5±0.5a 4.4±0.9a 5.8±1.2a 0±0a 3.8±0.2a 0±0a nt nt Pb (3840 μM) 3.5±0.4a 5.5±0.9a 6.4±1.4b 0±0a 4.1±0.3a 0±0a nt nt Pb (4800 μM) 4.8±1.6a 5.3±0.3a 5.8±1.1a 0±0a 5.1±0.4b 0±0a nt nt Pb (9600 μM) - 6.1±1a 3.8±0.5a 0±0a 4.6±0.4b 0±0a nt nt
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*Results are given as the mean and standard error of five (A. lappacea and A. harpophylla) or eight (T. australis and A. scabra) or either five or eight or ten (T. australis) replicates. ‘MGT R’ stands for ‘mean germination time of germinated seeds with radicle’; ‘MGT S’ stands for ‘mean germination time of germinated seeds with shoot only (without a radicle)’. Treatments with the same letter do not differ significantly (P>0.05) from the control. ‘SDIW’ stands for ‘sterilized deionized water’. ‘nt’ stands for ‘not tested’. ‘-‘: ‘no germinated seed’. ‘1st exp.’ refers to ‘first experiment’; ‘2nd exp.’ refers to ‘second experiment’.
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Appendix VIII. Shoot tolerance index (STI) of Astrebla lappacea, Themeda australis and Austrostipa scabra after 3 days on various concentrations of As(V) and Pb solutions. STI (%)* Metal/loid A. lappacea T. australis A. scabra Control (SDIW) 100±0a 100±0a 100±0a As(V) (13.34 μM) 107.2±15.7a 55.6±55.6a 181.9±76.1a AsV) (66.73 μM) 101.7±6.7a 6.7±6.7b 269.4±126.7a As(V) (133.47 μM) 113.8±16.2a 0±0b 286.1±118.8a As(V) (266.94 μM) 85.7±7.8a 20±20b 233.3±136.1a As(V) (400.41 μM) 66.2±18.2a 0±0b 98.6±38.8a As(V) (533.88 μM) 70.4±16.8a 0±0b 113.9±46.6a As(V) (667.36 μM) 65.8±31.5a 0±0b 204.2±103.9a Control (SDIW) 100±0a 100±0a 100±0a Pb (240 μM) 122.6±17.1a 50±50a 220.6±90a Pb (480 μM) 126.1±26.1a 4.2±4.2b 128.5±54.6a Pb (960 μM) 122.6±13.4a 0±0b 127.8±28.8a Pb (1920 μM) 101.6±6.1a 0±0b 137.8±41.9a Pb (3840 μM) 97.5±18.8a 0±0b 57.8±26.6a Pb (4800 μM) 104±21.5a 0±0b 33.8±24.6a Pb (9600 μM) 0±0b 0±0b 0±0a *Results are given as the mean and standard error of five (A. lappacea) or eight (A. scabra and T. australis) replicates. Treatments with the same letter do not differ significantly (P>0.05) from the control. ‘SDIW’ stands for ‘sterilized deionized water’.
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Appendix IX. Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings
The paper in this appendix was published as a research article in a high ranking refereed journal, Journal of Hazardous Materials, 2013 and in an international conference proceeding with following details:
Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whittaker, M., Pillai-McGarry, U., Schmidt, S. 2013. Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Journal of Hazardous Materials, Vol. 248-249, p.424-434.
Bigot, M., Guterres, J., Rossato, L., Pudmenzky, A., Doley, D., Whitakker, M., Pillai-McGarry, U., Schmidt, S. 2012. Novel metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings. Proceedings of the Seventh International Conference on Mine Closure 25-27 September 2012, Brisbane, Australia, Page 533-550.
For uniformity of the thesis, some changes in the formatting have been made including omitting the ‘article highlights’.
The experiment was conducted by the candidate and Ms Marie Bigot (the first author) with the assistance of Ms Christina Bee (honours student), Mrs Jenny Vu and Drs Laurence Rossato, Alex Pudmenzky, David Doley and Usha Pillai-McGarry. Production of metal-binding particles (X3) was funded by the University of New South Wales Centre for Advanced Macromolecular Design (CAMD). X3 particles were prepared by Dr Michael Whittaker. Soil psychrometer measurements and data interpretation were performed by Dr Usha Pillai-McGarry. Sample preparation and ICP analyses were conducted by the candidate and Ms Marie Bigot with the help of Ms Christina Bee and with the assistance of Mr David Appleton and Mr Stephen Appleton. Tailings and waste rock substrates were donated by operating and closed mine sites in Queensland (Australia).
The manuscript was written by Marie Bigot and reviewed and edited by the candidate with editorial advice from Drs Laurence Rossato, David Doley, Alex Pudmenzky, Michael Whittaker and Usha Pillai-McGarry.
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IX-1. Abstract Soil contaminants are potentially a major threat to human and ecosystem health and sustainable production of food and energy where mineral processing wastes are discharged into the environment. In extreme conditions, metal concentrations in wastes often exceed even the metal tolerance thresholds of metallophytes (metal-tolerant plants) and sites remain barren with high risks of contaminant leaching and dispersion into the environment via erosion. A novel soil amendment based on micron-size thiol functional cross-linked acrylamide polymer hydrogel particles (X3) binds toxic soluble metals irreversibly and significantly reduces their concentrations in the soil solution to below the phytotoxicity thresholds. X3 mixed into the top 50 mm of phytotoxic mine waste materials in pots in glasshouse conditions reduced total soluble concentrations of toxic contaminants by 90.3–98.7% in waste rock, and 88.6–96.4% in tailings, immediately after application. After 61 days, quality of unamended bottom layer of X3-treated pots was also significantly improved in both wastes. Combination of X3 and metallophytes was more efficient at improving soil solution quality than X3 alone. Addition of X3 to substrates increased substrate water retention and water availability to plants by up to 108% and 98% for waste rock and tailings, respectively. Soil quality improvement by X3 allowed successful early establishment of the native metallophyte grass Astrebla lappacea on both wastes where plants failed to establish otherwise.
IX-2. Introduction Toxic soil contaminants are a major threat to human and ecosystem health and sustainable food and energy production where up to 25,000 Mt of mineral processing wastes are discharged into the environment worldwide each year (Lottermoser, 2010). Soil acidity also increases the release of metals into the soil and their availability to organisms (Rensing and Maier, 2003). Accumulation of these metals in food chains can cause severe toxicity to both plants and animals, ultimately leading to highly degraded landscapes with bare soil prone to water and wind erosion, a major problem for global biogeochemical processes and human health and safety (Pimentel, 2006). Fewer than 5% of the 110,000 metal-contaminated sites in Australia have been remediated (CSIRO, 2004). Several established remediation techniques such as excavation and disposal to landfill, creation of physical barriers (i.e. cement, steel, bentonite, grout walls), or chemical amendments (i.e. red mud, zeolite, composts or lime) have been used to help detoxify, revegetate and stabilise landscapes (Garau et al., 2007, Khan and Jones, 2009, Mulligan et al., 2001, Bertocchi et al., 2006, Shi et al., 2009) thereby enhancing the future safety and environmental sustainability of disturbed areas (LPSDP, 2006). However, these remediation techniques are often very costly, environmentally disruptive and relatively ineffective, so areas may be left untreated.
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Metallophytes are plant species adapted to grow satisfactorily at extreme plant available soil metal concentrations without showing any sign of toxicity (Baker, 1981). The worldwide occurrence of metallophytes, both in natural and anthropogenic metal-polluted sites (Whiting et al., 2004), is a potential asset for cost-effective plant-based remediation strategies, or phytoremediation (Cunningham and Berti, 1993, Salt et al., 1995, Wu et al., 2010). This is especially true for phytostabilisation, by which metal-tolerant plants retain metals in the roots (Sinha et al., 2007) or immobilise contaminants through precipitation in the rhizosphere (Khan and Jones, 2009, Sinha et al., 2007). The potential for metallophytes to help rehabilitate highly contaminated lands has been studied extensively (Whiting et al., 2004, Salt et al., 1995, Baker, 1987, Lasat, 2000). Since the late 1980s, phytoremediation has been applied around the world to mine sites experiencing low to medium heavy metal contamination (Wong, 2003) but there has been limited application in Australia due to the few recognized native metallophyte species (Whiting et al., 2004, Pudmenzky et al., 2009). Exotic species are potentially invasive and therefore strictly prohibited for remediation of Australian soils for biosecurity reasons. Moreover, native species may be better adapted to local climatic conditions and low soil nutrient and salinity stresses (LPSDP, 2006, Mendez and Maier, 2008a). The establishment of pioneer metallophyte species is expected to improve soil quality (especially enhance soil water retention and reduce plant available metal concentration) and facilitate later colonisation by other less tolerant native shrubs and tree species able to stabilise the soil more effectively and extract water from deeper layers (Whiting et al., 2004).
The quality of plant growth media, fundamental for maintaining a healthy environment, is severely affected by two distinct waste streams at mine sites: waste rock that is discarded in the process of accessing the valuable ore, and tailings, which is the residue after extraction of minerals from the ore. Tailings are significantly finer grained (silt or sand size) than waste rock and are considered to be the major source of pollution at mine sites; in addition to containing heavy metals they carry reagents used for mineral extraction and their smaller particle size exposes more of the minerals to oxidative or reductive reactions (Mendez and Maier, 2008b). Soil quality is also often debased by additional high acidity and/or salinity, increasing the availability (and toxicity) of essential elements for plant growth and ecosystem functioning (ASEC, 2001). Unfortunately, high plant available metal concentrations resulting from mining often exceed the threshold for survival of even metallophytes (Wong, 2003, Mendez and Maier, 2008b) and substrate quality needs to be improved first before sowing or planting can be successful (Rossato et al., 2011).
To this end, Rossato et al. (2011) developed a novel micron-size thiol-functional cross-linked acrylamide polymer as a soil amendment (X3) to facilitate phytoremediation of phytotoxic soils and
263 plant growth media. The base material chosen for the synthesis of X3 is a hydrogel polyacrylamide (PAM), commonly applied in granular form to agricultural lands to improve soil structure, improve water infiltration during irrigation and reduce soil erosion (Rossato et al., 2011, Mendez and Maier, 2008b, ASEC, 2001). Not only can the functional thiols (–SH) attached to the particles bind irreversibly to metals, thereby effectively reducing metal bioavailability to plants in the soil (Rossato et al., 2011), but the high affinity of the hydrogel for water can also increase soil water holding capacity and storage after rainfall and thus provide water for plant uptake during drought (Rossato et al., 2011, Letey et al., 1992). Consequently, X3 can reduce both plant metal toxicity and water deficits, two major problems in Australian degraded lands. This technology, successfully tested under sterile laboratory conditions, showed promising results, especially in the promotion of healthy germination and early root elongation of two native Australian metallophyte grasses (Astrebla lappacea Lindl. Domin and Austrostipa scabra Lindl. S.W.L. Jacobs & J. Everett, Poaceae) under extremely inhospitable mine spoil conditions (high metal concentrations, salinity and low moisture conditions) (Guterres et al., 2013).
This study extended the findings of Rossato et al. (2011) and Guterres et al. (2013) and aimed to assess the effects of X3 metal-binding hydrogel polymer soil amendment alone or in combination with metallophyte grasses, under an extended range of environmental conditions on:
1. The capacity of X3 to ameliorate the poor plant growth media quality and physical characteristics of two contrasting toxic mine waste materials, mine waste rock and tailings.
2. The potential of X3 to promote healthy germination, emergence and early establishment of the native Australian metallophyte grass A. lappacea on both contaminated waste rock and tailings in a pot trial.
IX-3. Materials and methods IX-3.1. Materials IX-3.1.1. Substrate collection and characterisation Two mine spoils, waste rock and tailings, were obtained from Queensland, Australia, sites at which previous attempts at post-mining revegetation had failed. The waste rock was collected from an abandoned gold mine site in late 2009, to a depth of approximately 10 cm from the surface. Tailings were collected from a currently operating base metal mine site in February 2011, to a depth of approximately 20 cm from the surface. The control substrate was river sand, collected in April 2010 from Tamborine Mountain, Queensland, Australia and thoroughly washed with deionised water
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(DIW). All substrates were air dried, crushed, screened through a 2 mm sieve, mixed thoroughly in a soil mixer and stored in sealed containers at constant room temperature prior to experiments.
Chemical characteristics of the substrates were obtained by analysis for total metal concentration, using the hydrofluoric acid decomposition method (Loring and Rantala, 1992) and for plant available metal concentration, assessed after extraction with 0.01 M CaCl2 (Menzies et al., 2007) and measured using ICP-MS (inductively coupled plasma mass spectrometry) and ICP-OES (inductively coupled plasma optical emission spectroscopy). Soil pH and EC were determined using a digital EC and pH meter in 1:5soil:water (TPS 901-CP, TPS Pty Ltd.) after end-to-end shaking for 1 h (Rayment and Higginson, 1992).
IX-3.1.2. Application rate of X3 particles in the substrate mixtures Micron-size thiol-functional cross-linked acrylamide particles (X3) that were synthesised as described in Rossato et al. (2011) were mixed thoroughly with the waste rock and tailings substrates in a rotating drum for at least 12 h. The optimum X3 application rates for each mine substrate were determined based on plant available metal concentration and salinity levels of the substrates and were 38% and 19% of dry weight (DW) of the waste rock and tailings, respectively.
IX-3.1.3. Native seeds Viable seeds of the native metallophyte grass A. lappacea cv. Yanda were obtained from Native Seeds Pty Ltd (Cheltenham, Victoria, Australia). They had been harvested from Walgett (NSW) and stored for over 12 months to overcome dormancy.
IX-3.2. Glasshouse experiment A glasshouse experiment was conducted in physical containment level 2 (PC2) temperature controlled conditions set to 30˚C day and 25˚C night and with the addition of artificial light on a 12 h/12 h cycle. Pots (diameter 150 mm and height 150 mm) were lined with sealed plastic bags of 305 × 205 mm (BPM Trading Company Pty Ltd) and filled with the substrates to the rim of the pot. Seven treatments were applied in combination with seeds: unamended and X3-amended waste rock ([Waste Rock + 0% X3 (plants)] and [Waste Rock + 38% X3 (plants)]), unamended and X3- amended tailings [Tailings + 0% X3 (plants)] and [Tailings + 19% X3 (plants)], and the appropriate controls unamended sand [Sand + 0% X3 (plants)], sand amended with the amount of X3 corresponding to waste rock [Sand + 38% X3 (plants)] and sand amended with the amount of X3 corresponding to tailings [Sand + 19% X3 (plants)]. X3 amendment was applied to the surface (top 50 mm) of the pot while the bottom was filled with unamended material. Seeds were first surface sterilized using 20% bleach as described in Rossato et al. (2011), then sown at a density of 150
265 seeds per pot. Additional amended tailings [Tailings + 19% X3 (no plants)] and amended waste rock [Waste Rock + 38% X3 (no plants)] treatments were set up without plants in order to differentiate the effect of the particles and the metallophyte species on substrate condition. Treatments were replicated 3 times. Pots were watered daily by weight to maintain moisture levels at field capacity. Nutrient supply was non-limiting to avoid any plant nutrient deficiency that could affect plant growth during the trial and provided through three low strength nutrient solutions (A, B and C) specifically designed for Australian native species, each applied every three days. The −1 −1 −1 −1 solutions contained: (A) 0.0075 g L FeSO4, 0.03 g L MgSO4, 0.0625 g L KNO3, 0.00125 g L -1 -1 -1 CuSO4, 0.18 g L (NH2)2CO (urea); (B) 0.264 g L N (nitrate), 0.0614 g L N (ammonium), 0.1351 g L-1 N (urea), 0.1474 g L-1 P, 0.7398 g L-1 K (nitrate), 0.098 g L-1 K (sulphate), 0.042 g L-1 -1 -1 -1 -1 S, 0.0015 g L H3BO3, 0.0005 g L Zn (EDTA chelate), 0.0003 g L Mo, 0.0005 g L Cu (EDTA chelate), 0.0037 g L-1 Fe (DTPA chelate), 1.00 g L-1 Mn (EDTA) chelate); and (C) 3.00 g L-1
Ca(NO3)2.
IX-3.2.1. Determination of substrate quality IX-3.2.1.1. Plant available water and osmotic potential measurement Substrate water release characteristics in the presence and absence of X3 particles were assessed to estimate the amount of plant available water held between field capacity (−0.01 MPa) and permanent wilting point (−1.5 MPa) matric water potentials, ψm (Kramer and Boyer, 1995). Three treatments were tested for each of the substrates and included: (i) substrate (waste rock, tailings or sand) only (i.e. no X3 particles added), (ii) substrate + X3 applied at 19% of substrate DW and (iii) substrate + X3 applied at 38% of substrate DW. Each substrate and treatment had three replicates.
Additionally, due to the high salinity of the mine substrates, the osmotic potential, ψo, of the samples was derived by measuring the total water potential, ψT, on subsamples of all treatments and replicates following the matric potential measurements. Matric potential was determined using the pressure plate apparatus (Klute, 1986), wherein water-saturated substrate samples were subject to the appropriate air/nitrogen pressure in an enclosed chamber until water drainage from the samples ceased. Samples were removed from the apparatus and a subsample (∼30%) from each sample rapidly removed and stored in an airtight glass vial for total potential measurement, and the rest of the sample was oven-dried (105˚C) to determine gravimetric water content (Wc). Plant available water was calculated using Eq. (1):
PAW = Wc 0.01 − Wc 1.5 (1)
266 where PAW, plant available water (in % DW); Wc 0.01, percentage water content of the sample dry weight after 0.01 MPa pressure application; Wc 1.5, percentage water content of the sample dry weight after 1.5 MPa pressure application.
Total soil water potential for each subsample was measured soon after using a thermocouple psychrometer (SC-10, Decagon Devices, Pullman, Washington, USA) under constant temperature (23◦C) conditions.
Given that the components of ψT (Brady and Weil, 2004) measured under laboratory conditions are
ψT = ψm + ψo (2)
Osmotic potential was derived from Eq. (3):
ψo = ψT − ψm (3)
where ψo , ψT and ψm are expressed in −MPa.
IX-3.2.1.2. Redox potential measurement Soil redox potential (Eh) was measured immediately before seed sowing and 61 days after sowing to identify possible oxygen limitations associated with the addition of the X3 hydrogel (Bohn, 1971). A platinum electrode and an Ag/AgCl reference probe were connected to a voltmeter and inserted into the substrate at 2 cm depth. Measurements were taken after two minutes of equilibration and adjusted by +200 mV to convert to Eh (i.e. redox potential obtained when using a standard hydrogen electrode).
IX-3.2.1.3. Soil solution sampling and analysis The soil solution from both the top 50 mm and the bottom 100 mm layers of each pot was sampled using Rhizon Soil Moisture Samplers (EVH Engineering Pty Ltd.) immediately before sowing (day 0) and 61 days after sowing. Up to 10 mL of soil solution was extracted and 25 μL of 70% concentrated nitric acid was added per mL of soil solution to maintain metal solubility (Rossato et al., 2011). Samples were stored at 4˚C until metal analysis using ICP-MS and ICP-OES.
An additional 5 mL of soil solution was collected 61 days after sowing and analysed immediately for EC and pH using a calibrated digital EC and pH meter (TPS 901-CP, TPS Pty Ltd.).
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IX-3.2.2. Substrate penetration resistance Substrate penetration resistance was measured 27 days after sowing to test the effect of X3 on the surface strength of extremely saline substrates. Three readings were made using a Geotester pocket penetrometer with a 15 mm tip (Facchini, Italy) that was inserted into the surface of the substrate in each pot to a depth of 10 mm and the readings averaged per pot (this constituted one replicate per treatment). As soil penetration resistance is governed by soil water content (Taylor and Brar, 1991) or potential, readings were made 24 h after pots were watered to field capacity (−0.01 MPa) water content in an effort to reduce effects of water potential difference between treatments and expressed in kPa.
IX-3.2.3. Determination of emergence, plant establishment and shoot height Seedling emergence per pot was recorded daily for the first 13 days after A. lappacea seeds were sown. Plant establishment, survival and maximum shoot height were monitored weekly from day 13 onwards. Maximum height was determined by averaging the shoot height of the five tallest plants from each pot (from the base of the shoot to the tip of the longest leaf). Plants were also inspected daily for visible symptoms of metal toxicity and/or nutrient deficiencies.
IX-3.3. Statistical analysis Data were presented as means and standard errors of 3 replicates, and they were tested for significant differences between treatment means using Student’s t-tests (and a 95% confidence interval) using Microsoft Excel 2010. Statistical significance was indicated by P-values <0.05.
IX-4. Results IX-4.1. Substrate characteristics Waste rock was very acidic (pH = 3.10) with high plant availabilities of aluminium (Al), copper (Cu), zinc (Zn), manganese (Mn) and cobalt (Co) (Table IX-1), which were above the phytotoxicity thresholds for these metals (Mulvey and Elliott, 2000). Tailings, on the contrary, had neutral pH (7.03), therefore lower metal plant availabilities, although Zn, Mn, Co, and cadmium (Cd) concentrations were elevated (Table IX-1) and greater than their respective phytotoxicity thresholds (Mulvey and Elliott, 2000). ECs of 2.87 and 5.54 dS m-1 for waste rock and tailings (respectively) (Table IX-1) indicate that both substrates were extremely saline (Mulvey and Elliott, 2000, The State of Queensland, 2011). Sand had no metal contamination and was therefore an appropriate control substrate (Table IX-1).
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Table IX-1. Chemical characteristics of waste rock, tailings and sand substrates. Plant -1 available metal concentration (mg kg ) was assessed via 0.01 M CaCl2 extraction. Data are presented as the mean ± standard error (SE) (n=3).
Waste rock Tailings Sand (control)
Concentration of key metal contaminants (mg kg-1)
Element Total Plant available Total Plant available Total Plant available
Al 33,395983 621.6±3.7 10,546±164 1.78±1.19 15,049±294 0.00±0.00 Cu 1690±27 72.31±0.43 5002±75 1.49±0.08 0.00±0.00 0.00±0.00 Zn 309.3±6.2 10.30±0.06 6617±24 24.04±0.32 7.98±0.73 0.00±0.00 Mn 375.2±12.4 22.32±0.10 1733±7 32.73±0.15 167.4±1.5 0.18±0.03 Co 11.03±0.59 0.43±0.05 372.0±3.6 3.62±0.02 2.19±0.18 0.05±0.01 Cd 3.67±0.42 0.00±0.00 29.06±1.80 1.32±0.02 2.27±0.21 0.00±0.00
Other characteristics
Waste rock Tailings Sand (control)
-1 EC1:5water (dS m ) 2.87±0.03 5.54±0.02 0.02±0.00 pH 3.10±0.01 7.03±0.00 6.71±0.11 1:5water
IX-4.2. Capacity of X3 to improve substrate quality and penetration resistance IX-4.2.1. Effect of X3 on plant available water and osmotic potential Results from the pressure plate apparatus showed that the addition of X3 significantly increased (P<0.05) not only the water holding capacity of all substrates but also the plant available water (PAW) by up to 108% and 98% for waste rock and tailings, respectively, for the greater X3 concentration (38% DW). A linear trend of increasing PAW with increasing X3 application rates is apparent within the scope of the experiment (Figures IX-1A and IX-1B).
In the case of extremely saline growth media such as waste rock and tailings (Table IX-1) (The State of Queensland, 2011) osmotic potential can have a considerable effect on total water potential, and consequently PAW, due to the low chemical potential of the water, making water uptake difficult. Hence, PAW derived from the measured matric potential (−0.01 MPa and −1.5 MPa) water contents (Figures IX-1A and IX-1B) will overestimate this parameter (from Eq. (2)) if osmotic potential is less than zero.
Figures IX-1C and IX-1D show that, with the addition of X3 particles to the waste rock and tailings, osmotic potential was significantly increased compared to treatments without particles, especially at
269 the larger (38% DW) concentration of X3 for the waste rock where it was close to zero. This suggests that, except for the 38% X3 treatment, the PAW obtained for the other waste rock treatments would be overestimates, with the greatest overestimation being for the waste rock only (0% X3) treatment.
For the tailings (Figure IX-1D), the addition of X3 increased osmotic potential and this was significant at field capacity only (P<0.05). The amount of added X3 had no significant effect (P>0.05) on osmotic potential at the permanent wilting point, although an increasing trend of osmotic potential with the addition of X3 was evident.
Figure IX-1. Water characteristics of substrates in the presence of X3: plant available water in waste rock treatments and corresponding sand controls (A) and tailings treatments and corresponding sand controls (B); and osmotic potential of waste rock treatments (C) and tailings treatments (D) at field capacity (−0.01 MPa) and permanent wilting point (−1.5 MPa). Data are represented as the mean ± standard error (SE) (n=3). Asterisks indicate data based on 1 replicate only. Treatments with the same letter do not differ significantly (P>0.05).
IX-4.2.2. Effect of X3 on substrate redox potential On the basis that a biological environment is anaerobic between a redox potential (Eh) of −250 and +350 mV (Delaune et al., 1990), our study indicated that all waste rock substrates were aerobic with
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Eh significantly (P<0.05) higher than the +350 mV threshold (Figure IX-2A). The presence of plants did not affect Eh of X3-amended waste rock where it remained constant (approximately +550 mV) over time (P>0.05, Figure IX-2A). Initial anaerobic conditions were observed in both X3- amended sand substrates (+291.8±1.9 mV [Sand + 38% X3], Figure IX-2A; +272.9±3.0 mV [Sand + 19% X3], Figure IX-2B), and X3-amended tailings had Eh just slightly below the limit for aerobic conditions (+342.6±1.2 mV [without plant], +344.0±1.0 mV [with plants], Figure IX-2B). While aerobic conditions were restored over time for both sand treatments, X3-amended tailings (+307.23±12.77 mV [no plants], +342.6±11.48 mV [plants], Figure IX-2B) remained slightly below the aerobic threshold for the duration of the trial. However, the tailings treatments with X3 but no plants showed significantly reduced Eh over time (P=0.05, Figure IX-2B), while pots with plants remained unchanged over time (P>0.05, Figure IX-2B).
Figure IX-2. Redox potentials of substrates in pots before sowing (white bars) and 61 days after sowing (dark bars) for waste rock treatments and corresponding controls (A) and tailings treatments and corresponding controls (B). Data are represented as the mean ± standard error (SE) (n=3). Treatments with the same letter do not differ significantly (P>0.05).
IX-4.2.3. Effect of X3 on soil solution quality With the addition of X3 to the top layer of the pots only and no plants, the substrate solution ECs of both top and bottom layers of the two mine waste materials decreased significantly (P<0.05) after 61 days (Table IX-2). The presence of plants on X3-amended pots further reduced EC in both layers of waste rock and tailings (P<0.05, Table IX-2). When X3 and plants were combined, pH of both top and bottom layers of waste rock increased significantly (P<0.05) from 2.54±0.05 to 3.84±0.20 and from 2.33±0.00 to 2.86±0.02, respectively (Table IX-2) 61 days after sowing, and this increase of pH in both layers of waste rock was greater than when X3 was applied without plants (Table IX- 271
2). Similar pH trends were observed in both layers of tailings (P<0.05) but not in control sand (Table IX-2).
Table IX-2. Electrical conductivity (EC, dS m-1) and pH of soil solution from each layer of mine substrates 61 days after sowing. Data are represented as the mean ± standard error (SE) (n=3).
Waste rock
Treatment 0% X3 (plants) 38% X3 (no plants) 38% X3 (plants)
Pot layer Top Bottom Top Bottom Top Bottom
EC Mine substrate 12.44±1.23 13.01±0.91 6.24±0.89 9.82±0.37 3.51±0.51 4.31±0.09
Control (sand) 0.02±0.00 0.09±0.01 - - 0.15±0.04 0.13±0.01 pH Mine substrate 2.55±0.05 2.33±0.00 3.73±0.18 2.56±0.02 3.84±0.20 2.86±0.02
Control (sand) 5.91±0.08 6.74±0.04 - - 5.48±0.24 6.77±0.13 Tailings
Treatment 0% (plants) 19% (no plants) 19% (plants)
Pot layer Top Bottom Top Bottom Top Bottom
EC Mine substrate 80.77±1.40 62.07±5.75 19.87±1.13 32.66±1.99 13.93±1.29 16.75±1.60
Control (sand) 0.02±0.00 0.09±0.01 - - 1.30±0.88 0.33±0.01 pH Mine substrate 7.07±0.04 7.18±0.03 7.58±0.03 7.44±0.05 7.73±0.02 7.52±0.02
Control (sand) 5.91±0.08 6.74±0.04 - - 5.69±0.18 7.16±0.10
Although the chemical compositions of the two waste materials were very different, the addition of X3 drastically reduced plant availability of all major soluble toxic substrate contaminants, immediately in the top layer and over time in the unamended bottom layer (Tables IX-3 and IX-4). In waste rock, addition of X3 rapidly reduced Al, Cu, Zn, Co and Mn concentrations by 98.7, 98.7, 94.5, 90.3 and 91.3% at day 0 (Table IX-3) in the top layer and metal concentrations were maintained below the phytotoxicity threshold for A. lappacea throughout the duration of the experiment. In the bottom layer of the amended waste rock treatment, metal concentrations were gradually decreased by 96.5, 94.4, 95.4, 95.6 and 91.2% (respectively) after 61 days (Table IX-3). In tailings, similar trends were observed and reductions of Mn, Zn, Co and Cd concentrations by 94.7, 96.4, 92.4 and 89.7% (respectively) occurred immediately in the top layer of the X3-amended treatments compared to the unamended tailings (Table IX-4). The concentrations of these metals in the bottom layer of the amended tailings treatments were reduced gradually by 68, 84.1, 64.9 and 82.6% (respectively) after 61 days (Table IX-4). S and Na concentrations also decreased
272 significantly by 88.6 and 89% (respectively) in tailings immediately after X3 was applied (from 41,177±3617 mg L-1 and 2829.8±134.3 mg L-1 in unamended tailings to 4708±554 mg L-1 and 307.4±39.3 mg L-1 in the top layer of the amended tailings, Table IX-4). Reductions of S and Na concentrations in the amended top layer were maintained over time, with no significant difference between solution concentrations at day 0 and day 61 (P>0.05, Table IX-4). In the bottom layer of the amended tailings, S and Na concentrations were gradually reduced by 74 and 69% after 61 days compared to unamended pots (Table IX-4). Overall, in both layers of waste rock and tailings, the combination of X3 and metallophyte grass was more effective at improving soil solution quality than X3 alone (P<0.05) (Tables IX-3 and IX-4).
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Table IX-3. Concentration of key contaminants in soil solution (mg L-1) from each layer of waste rock and control sand treatments before sowing (Day 0) and 61 days after sowing. Data are represented as the mean ± standard error (SE) (n=3).
Element Time Day 0 Day 61
Treatment 0% X3 (plants) 38% X3 (no plants) 38% X3 (plants) 0% X3 (plants) 38% X3 (no plants) 38% X3 (plants)
Pot layer Top Bottom Top Bottom Top Bottom Top Bottom Top Bottom Top Bottom
Al (mg L-1) Waste rock 1415±57 1647±153 14.81±3.05 1456±62 18.9±4.79 1175±22 1108±208 1385±163 69.02±40.67 563.5±252.2 8.88±3.53 48.79±4.47
Sand 0.19±0.02 0.16±0.03 - - 1.16±0.19 0.15±0.01 0.1±0.08 0.14±0.05 - - 0.55±0.24 0.06±0.02
Cu (mg L-1) Waste rock 152.0±8.8 176.0±15.7 1.59±0.25 155.9±6.4 1.99±0.43 132.1±2.6 97.53±15.43 133.4±12.4 7.68±2.77 53.02±21.30 1.69±0.59 7.43±0.67
Sand 0.02±0.00 0.01±0.00 - - 0.06±0.02 0.01±0.00 0.01±0.00 0.01±0.00 - - 0.16±0.08 0.02±0.00
Zn (mg L-1) Waste rock 23.71±0.95 27.51±2.32 1.21±0.17 24.01±1.00 1.30±0.26 19.71±0.30 17.42±2.49 19.82±1.77 1.50±0.77 8.39±3.56 0.31±0.08 0.91±0.07
Sand 0.02±0.00 0.02±0.00 - - 0.04±0.01 0.01±0.00 0.02±0.00 0.03±0.00 - - 0.13±0.02 0.08±0.03
Mn (mg L-1) Waste rock 42.70±1.44 49.42±3.63 3.68±0.31 43.63±0.92 3.72±0.55 35.83±0.91 241.8±15.7 84.53±3.31 12.37±4.12 21.67±7.39 16.28±2.11 7.47±0.42
Sand 0.04±0.00 0.03±0.00 - - 0.04±0.01 0.03±0.01 0.69±0.38 0.01±0.00 - - 1.93±0.69 0.09±0.05
Co (mg L-1) Waste rock 1.52±0.05 1.73±0.12 0.16±0.01 1.55±0.07 0.15±0.03 1.29±0.03 1.27±0.17 1.37±0.12 0.12±0.06 0.60±0.26 0.04±0.01 0.06±0.01
Sand 0.03±0.00 0.02±0.00 - - 0.03±0.00 0.05±0.01 0.03±0.00 0.01±0.00 - - 0.26±0.09 0.04±0.01
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Table IX-4. Concentration of key contaminants in soil solution (mg L-1) from each layer of tailings and control sand treatments before sowing (Day 0) and 61 days after sowing. Data are represented as the mean ± standard error (SE) (n=3).
Element Time Day 0 Day 61
Treatment 0% X3 (plants) 19% X3 (no plants) 19% X3 (plants) 0% X3 (plants) 19% X3 (no plants) 19% X3 (plants)
Pot layer Top Bottom Top Bottom Top Bottom Top Bottom Top Bottom Top Bottom
Zn (mg L-1) Tailings 67.77±4.09 59.16±1.67 3.2±0.31 54.18±0.64 2.47±0.29 52.85±1.27 33.88±2.48 57.05±5.11 11.17±1.07 18.51±2.25 5.88±0.36 9.08±0.62
Sand 0.02±0.00 0.02±0.00 - - 0.04±0.01 0.05±0.01 0.02±0.00 0.03±0.00 - - 0.21±0.04 0.28±0.15
Mn (mg L-1) Tailings 90.99±5.02 79.42±1.99 6.93±0.67 71.68±2.06 4.82±0.49 70.56±0.87 501.0±4.2 200.1±108.3 55.11±4.40 119.0±6.1 24.56±2.62 63.49±11.62
Sand 0.04±0.00 0.03±0.00 - - 0.07±0.03 0.08±0.01 0.69±0.38 0.01±0.00 - - 4.01±1.23 1.79±1.23
Co (mg L-1) Tailings 6.82±0.36 6.20±0.14 0.67±0.02 5.72±0.08 0.52±0.04 5.59±0.07 5.50±0.30 6.44±0.31 3.02±0.25 4.00±0.34 1.80±0.07 2.26±0.12
Sand 0.03±0.00 0.02±0.00 - - 0.04±0.00 0.03±0.00 0.03±0.00 0.01±0.00 - - 0.23±0.01 0.23±0.11
Cd (mg L-1) Tailings 1.71±0.15 1.45±0.05 0.21±0.02 1.38±0.02 0.18±0.01 1.29±0.03 0.73±0.06 0.62±0.04 0.10±0.01 0.16±0.01 0.05±0.00 0.11±0.01
Sand 0.03±0.00 0.05±0.01 - - 0.04±0.01 0.02±0.00 0.01±0.00 0.01±0.00 - - 0.01±0.00 0.01±0.00
Na (mg L-1) Tailings 2830±134 2330±62 392.6±20 2100±45 307.4±39.3 2083±24 2024±79 1691±144 754.0±87.7 911.3±12.8 415.3±61.9 510.4±67.2
Sand 9.43±0.19 11.99±0.58 - - 2.06±0.04 10.25±0.52 0.35±0.08 3.21±0.24 - - 14.51±1.99 13.38±6.10
S (mg L-1) Tailings 41,177±3617 32,615±1205 5841±321 30,638±402 4708±554 30,000±432 26,863±780 20,416±2475 5916±394 10,815±1004 3484±501 5125±804
Sand 8.26±1.30 7.25±1.50 - - 13.44±0.88 9.20±1.06 0.35±0.05 0.35±0.02 - - 5.57±1.51 3.05±1.80
275
IX-4.2.4. Effect of X3 on substrate penetration resistance In the absence of X3, all substrates had significantly greater (P<0.05) surface penetration resistance compared to the treatments with X3 (Figure IX-3). Values were 200±11 kPa for unamended sand, 488±42 kPa for unamended waste rock, and 856±74 kPa for unamended tailings during plant establishment (Figure IX-3). Apart from the sand, the mine substrate values were large enough to potentially restrict root penetration and exploration (Haling et al., 2011). On the contrary, substrate penetration resistance was substantially (P<0.05) lower in all substrates treated with X3 (Figure IX- 3) suggesting that they were no longer detrimental to root penetration (Haling et al., 2011).
Figure IX-3. Penetration resistance of X3-amended substrates and corresponding controls 27 days after sowing for waste rock treatments and corresponding controls (A), and tailings treatments and corresponding controls (B). Data are represented as the mean ± standard error (SE) (n=3). Treatments with the same letter do not differ significantly (P>0.05).
IX-4.3. Effect of X3 on seedling emergence, plant establishment and growth Emergence of A. lappacea seeds was greatly promoted by X3 on both mine waste materials (Figures IX-4 and IX-5). With X3, seedling emergence percentage was significantly increased from 0% on both unamended substrates to 62% on waste rock (Figure IX-4A) and 54% on tailings (Figure IX-5A). All pots had reached maximum emergence after 13 days.
Twenty days after sowing, no emergent seedling had survived on any of the unamended spoils (Figures IX-4 and IX-5). In contrast, early establishment of A. lappacea on all other treatments was successful (Figures IX-4B and IX-5B). Survival reached a plateau after day 34, with no further significant plant death in both X3-amended mine substrates (Figures IX-4B and IX-5B). In X3- amended waste rock, plant establishment percentage (45%, Figure IX-4B) was not significantly
276 different from the amended sand control (41%, Figure IX-4B). Although X3-amended tailings had lower establishment percentages (35%, Figure IX-5B) than the amended sand control (61%, Figure IX-5B), plants had established at a satisfactory percentage when compared to unamended tailings treatments (0%, Figure IX-5B).
Maximum shoot height of A. lappacea on waste rock amended with X3 significantly increased (P<0.05) to up to 28.3 cm after 55 days (Figure IX-4C) but remained relatively lower (P<0.05) compared to the two sand controls (above 44 cm, Figure IX-4C). On tailings amended with X3, maximum shoot height was also significantly (P<0.05) lower than controls (Figure IX-5C). However, in both amended mine substrates, successfully established plants were vigorous and healthy and did not display any visible signs of toxicity on leaves throughout the experiment (Figures IX-4D and IX-5D).
Figure IX-4. Development of Astrebla lappacea in pots with waste rock and corresponding controls. Seedling emergence (A), plant establishment (B), maximum plant height (C), and photograph showing waste rock pots unamended (left) and amended with X3 particles (right) 20 days after sowing (D). Vertical bars when larger than the symbol indicate ± SE of the mean for n=3. The wide SE observed in (A) and (B) are due to the Sand + 38% X3 treatment with one replicate displaying low emergence and establishment percentages compared to the other two.
277
Figure IX-5. Development of Astrebla lappacea in pots with tailings and corresponding controls. Seedling emergence (A), plant establishment (B), maximum plant height (C), and photograph showing tailings pots unamended (left) and amended with X3 particles (right) 20 days after sowing (D). Vertical bars when larger than the symbol indicate ± SE of the mean for n=3.
IX-5. Discussion This study provides a more detailed understanding of the novel X3 remediation technique in terms of interactions between X3 particles and two contrasting toxic metalliferous mine waste materials: highly acidic, extremely saline and metal-polluted waste rock, and neutral extremely saline tailings with lower metal availability to plants. With the addition of X3, substrate solution quality was immediately improved in both waste rock and tailings by marked reductions in contaminant metal concentrations below phytotoxicity thresholds in both waste rock (Al, Cu, Zn, Mn and Co) and tailings (Zn, Mn, Co and Cd). Our study showed that X3 particles were very versatile and were capable of adapting to various types of contaminations in the differing substrates; i.e. X3 had high affinities for soluble Al, Cu and Zn in waste rock (94.5–98.7% reduction in concentration), whereas soluble Zn was the metal most extensively sequestered in tailings (96.4%). The increase in pH for X3-amended substrates (both waste rock and tailings) was an additional factor that may have slightly reduced metal solubility (Sauvé et al., 2000). Decreased metal concentrations in 278 contaminated soils has been demonstrated using polyacrylate, another hydrogel polymer, particularly reductions of lead (Pb) by up to 66% in Pb-contaminated spoil (Guiwei et al., 2008) and reductions of Cd, nickel (Ni) and Zn (together in a same substrate) by up to 47% (de Varennes et al., 2006). However, the polyacrylates used by Guiwei et al. (2008) and de Varennes et al. (2009) reduced soil urease activity, and the metal-binding and water-holding effects were competitive. In contrast, X3 was highly efficient at both reducing various metals present at phytotoxic plant available concentrations (by up to 98.7%) and increasing plant available water levels for plants which confirms the potential of this technology for in situ remediation of mine spoils, particularly under arid and semi-arid environments.
Our study of X3 particles showed that this synthetic polymer increases water-holding capacity in both waste rock and tailings, while its hydrogel structure does not impose substrate oxygen limitations. X3 water-holding capacity was not hampered by the extreme salinity of these substrates in solution as opposed to some other gel-forming synthetic polymers (Johnson, 1984). Our findings are consistent with Shi et al. (2010) who demonstrated that the hydrophilic properties of a polyacrylamide-based polymer (Stockosorb) excluded Na+ and Cl− from the substrate while the high water holding capacity of the polymer effectively provided desalinated water to the plant. The osmotic potential of the waste rock and tailings on their own were sufficiently low (-2 MPa) to hinder plant growth in these media without amendment. The addition of X3 particles raised osmotic potential in both substrates and thereby contributed to greater PAW by increasing both osmotic potential and water holding capacity, and could also contribute to lowering the effect of salt toxicity on plant growth. Plant water availability and osmotic potential were lower in X3-amended tailings compared to X3-amended waste rock. Factors that may have influenced the difference between the substrates include (a) the greater (almost twice) EC1:5water of the tailings compared to the waste rock, (b) the nature of the salts in each substrate, as different salts were shown to impact polyacrylamide hydrogels differently (Johnson, 1984), and (c) substrate particle size, where the tailings were finer-grained than waste rock and thus would retain a greater volume of solution and therefore a greater mass of dissolved salts at −1.5 MPa ψm. X3 considerably reduced salinity of the substrates, especially for tailings, where EC of unamended substrate solution was higher than the EC for seawater (>51 dS m−1) (Suttar, 1990). The concentrations of non-metallic substrate contaminants in solution, particularly S and Na in tailings, were reaching highly toxic concentrations (Mulvey and Elliott, 2000) and our study showed that X3 was also capable of greatly reducing their concentrations in solution, by up to 89%. X3 particles increased the osmotic potential and as a consequence improved PAW while providing oxygenated conditions in both the waste rock and tailings, both of which were extremely saline media. Increase in the osmotic potential indicates
279 a reduction in the salt concentration and a more desirable plant growth environment. Our findings are consistent with other studies on the capacity of superabsorbent polymers to reduce salinity and improve plant survival (by increasing water-use efficiency under osmotic stress) (Dorraji et al., 2010, Hüttermann et al., 2009, Sayed et al., 1991).
Values of penetration resistance measured in the surface of the unamended waste rock and tailings were found to affect root penetration of perennial grass species (Campbell and Swain, 1973), particularly in acidic conditions (Haling et al., 2011). The high values for the unamended mine substrates were obtained despite the high water content (close to field capacity) during measurement, suggesting that further drying of the substrate surface would increase its strength to levels that would prevent seedling emergence and root growth. The addition of the X3 hydrogel significantly modified the physical characteristics of both mine materials and reduced substrate penetration resistance providing a suitable environment for emergence, root penetration and exploration and plant establishment (Haling et al., 2011, Guiwei et al., 2008).
Our study demonstrated that the X3 remediation technique successfully alleviated phytotoxicity in highly contaminated mine waste rock and tailings where plant establishment had previously failed. As a result, X3 promoted successful emergence and early establishment of the native metallophyte species A. lappacea on both mine wastes. In contrast with other nanoparticles such as “Fe3O4 or
TiO2 nanopowders” or other gel-forming polyacrylamide soil conditioners which were found to, respectively, reduce germination potential of cucumber seeds (Mushtaq, 2011), or inhibit germination of lettuce (Woodhouse and Johnson, 1991), our X3 particles were not toxic to seed germination, emergence or plant establishment. The capacity of X3 to mitigate the effects of different types of contamination such as metals, salinity, acidity and substrate penetration resistance confirms the technology as an asset for mine waste remediation. Another polyacrylate polymer was previously shown to reduce metal concentrations in contaminated substrates and to allow the establishment of perennial ryegrass (de Varennes et al., 2006, de Varennes et al., 2009, de Varennes and Torres, 1999), but this polymer effectively reduced soluble Cu in substrate in the presence of plants (ryegrass) only (de Varennes and Torres, 1999). In contrast, X3 polymer significantly reduced Cu and other metal contamination whether it was applied with plants or without. The combination of X3 and the metallophyte grass A. lappacea was, however, shown to be more beneficial to substrate quality than X3 alone in terms of pH, EC, and metal and other contaminant concentrations in substrate solution. Although further investigations are needed to understand the effects of X3 particles on soil-plant relationships and its applicability in the field, the X3 remediation technology shows great promise as an effective amendment in tandem with
280 phytoremediation to maximise the rehabilitation potential of a site in a sustainable manner. The cost of the X3 technology will depend on the level of contamination at sites and method of application employed, but should be in the range of other polyacrylamide-based soil conditioners which recently became widely applied for degraded land management because they were shown to be sustainable, environmentally friendly and highly efficient and inexpensive (Guiwei et al., 2008, Hüttermann et al., 2009, Orts et al., 1999, Sojka et al., 2007), with cost estimated at US$ 2–4 kg−1 (Hüttermann et al., 2009) or € 2 kg−1 (Guiwei et al., 2008).
IX-6. Conclusion This study showed that X3 particle amendment of surface mine substrates alleviated metal and salinity toxicity, improved substrate penetration resistance, acidity and plant water relations and allowed germination and healthy establishment of the native metallophyte grass A. lappacea in the polluted mine waste rock and tailings tested. This novel remediation technique promises to be very robust, cost-effective and applicable to a wide range of situations where decades of attempted plant establishment had previously been unsuccessful. Further study is required to test the sustainability of the X3 technology via a long-term glasshouse experiment to determine whether the plant establishment enabled by X3 can be sustained. Future research will also examine the potential toxicity of X3 particles to soil biota, and the X3 technique (including its potential longevity, methods of particle deployment and the effects of different seasonal conditions on particle properties and plant growth as well as interaction of the particles with soil microbes) will be tested and validated in a pilot field trial at a mine site under extreme environmental conditions.
IX-7. Acknowledgements This study was supported by Xstrata Technology, Commercialisation Australia and the Department of Employment, Economic Development and Innovation (DEEDI Mines). The assistance of mine site personnel is gratefully acknowledged.
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Appendix X. A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters.
The appendix consists of a hyperlink to a journal article that has been published in the high ranking refereed journal, Journal of Hazardous Materials 2019, vol. 364, 449-467, under the title: ‘A new conceptual framework for plant responses to soil metals based on metal transporter kinetic parameters’. https://doi.org/10.1016/j.jhazmat.2018.09.026
Appendix XI. Assessing germination characteristics of Australian native plant species in metal/loid solution.
The appendix consists of a hyperlink to a journal article that has been published in the high ranking refereed journal, Journal of Hazardous Materials 2019, vol. 364C, 173-181., under the title: ‘Assessing germination characteristics of Australian native plant species in metal/metalloid solution’. https://doi.org/10.1016/j.jhazmat.2018.10.019
Appendix XII. Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings.
The appendix consists of a hyperlink to a journal article that has been published in the high ranking refereed journal, Journal of Hazardous Materials 2013, vol. 248-249, 442-450, under the title: ‘Micron-size metal-binding hydrogel particles improve germination and radicle elongation of Australian metallophyte grasses in mine waste rock and tailings’. https://ac-els-cdn-com.ezproxy.library.uq.edu.au/S0304389413000678/1-s2.0- S0304389413000678-main.pdf?_tid=f5fc8028-770a-494f-80bb- 1ee6189cb885&acdnat=1522305103_14cb50b9248228527f2a7f4fe6da403a
Appendix XIII. Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings.
The appendix consists of a hyperlink to a journal article that has been published in the high ranking refereed journal, Journal of Hazardous Materials 2013, vol. 248-249, 424-434, under the title: ‘Metal-binding hydrogel particles alleviate soil toxicity and facilitate healthy plant establishment of the native metallophyte grass Astrebla lappacea in mine waste rock and tailings’.
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