Effects of magnetite removal on the distribution and ...341651/s... · Effects of magnetite removal...

Effects of magnetite removal on the distribution and speciation of Arsenic in

copper tailings and its accumulation in native grass

Yunjia Liu

Bachelor of Science

A thesis submitted for the degree of Master of Philosophy at

The University of Queensland in 2014

Sustainable Minerals Institute

Abstract

Arsenic (As) is considered a significant pollutant in neutral-alkaline copper (Cu)

tailings, which may be transported off site via seepage and/or runoff from tailings

storage facilities. Iron (Fe) oxides and oxyhydroxides, with their high specific surface

area and As affinity, play an important role adsorbing inorganic As in contaminated

soil and water. In addition, organic matter (OM) and the resulting organic molecules

can not only directly influence As chemical forms through competing functional

groups of OM (such as phenolic, carboxyl, hydroxyls, etc.), but OM can also catalyse

the transformation of Fe-primary minerals into secondary minerals such as Fe-

oxyhydroxides via microbe-mediated processes. At the Ernest Henry Mine (North

Queensland, EHM), the Cu ore processing circuit has recently been modified to

recover magnetite (Fe3O4) from tailings, which could reduce magnetite concentration

in the tailings from 20–30% to as low as 3-5%. However the environmental risks of

As mobility in the low magnetite (LM) tailings are yet to be investigated.

The present study aimed to investigate the effects of magnetite removal and direct

revegetation treatments on As distribution, solubility, and speciation in relation to

plant As accumulation in the Cu-tailings. It is hypothesized that the distribution of As

into exchangeable and soluble forms may be increased in the low magnetite (LM)

tailings, due to the much reduced As adsorption capacity associated with the

magnetite and the resultant Fe-oxyhydroxides coating on the surfaces of magnetite

following redox processes. In addition, the increased distribution of As into the pore

water of the LM tailings may favour As conversion from the inorganic into the organic

forms under organic matter amendment and direct revegetation with native grass

species. Both LM and high magnetite (HM) tailings were amended with 5%

sugarcane residues as a basal treatment to ensure plant survival, in combination

with 0, 1 and 5% pine-biochar, in which native grass Red Flinders (Iseilema

Vaginiflorum) plants were grown for 4 weeks under glasshouse conditions. Arsenic

distribution in the LM and HM tailings was fractionated before and after direct

revegetation treatments.

The intrinsic As adsorption capacity in the HM tailings was significantly higher than

that in the LM. Following the organic matter and direct revegetation treatments, As

distribution in the specifically adsorbed and amorphous Fe oxyhydroxide increased

in the LM tailings but declined in the HM tailings, compared to the unamended

tailings. Total As concentration in the pore water of LM tailings increased by 6-7 fold

compared to those in the HM tailings. In the amended and revegetated tailings

treatments, As concentrations in the pore water of the LM tailings were significantly

elevated compared to those in the HM. The proportion of inorganic As species in the

pore water of the LM tailings was approximately 50% higher than those in the HM.

After 4 weeks of treatment, the native grass accumulated up to 137 mg kg-1 As in the

roots and 2 mg kg-1 As in the shoots. Based on current findings and literature review

on Cu-tailings, As adsorption capacity in the tailings decreased with lowering

magnetite contents. This may be attributed to the combined adsorption effects of

magnetite itself and the newly formed Fe-oxyhydroxides coating the magnetite

particle surfaces, resulting in increased As distribution into the pore water. Under

amendment and direct revegetation, organic matter added in the tailings may have

stimulated microbial mediated Fe dissolution and formation of amorphous Fe-

oxyhydroxides at the surfaces of magnetite particles. This therefore resulted in

reduced As uptake in the native grass grown in the HM tailings.

The findings have demonstrated the regulatory roles of Fe-minerals such as

magnetite on As mobilisation and transformation in the neutral-alkaline Cu-tailings

under revegetation. The findings further highlight the potential impacts of ore

processing on tailings and the risk of environmental pollution. Further studies are

required to illustrate the detailed mechanisms of mineral transformation and roles of

key microbial processes at the interface of mineral particles and roots of diverse

native plants species.

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, 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 research higher degree 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 General Award Rules of The University of

Queensland, immediately made available for research and study in accordance with

the Copyright Act 1968.

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.

Publications during candidature

Conference Paper

YJ Liu, L Zhao, L Huang (2014) Arsenic bioavailability regulated by Magnetite in

Copper Tailings: As mobilization into pore water and plant uptake, in: 5th

International Congress on Arsenic in the Environment. Buenos Aires, Argentina.

Publications included in this thesis

No publications included

Contributions by others to the thesis

No contributions by others

Statement of parts of the thesis submitted to qualify for the award of another

degree

None

Acknowledgements

I am greatly indebted to my parents for their tremendous help financially and

emotionally. Their unconditional love enables me to understand the value of life.

Without their help, I would have had the opportunity to undertake my MPhil study at

the University of Queensland.

I would like to express my sincere thanks to my principal supervisor Dr Longbin

Huang. He is the person who always shares his successful experiences with me. He

has provided his consistent guidance and academic support throughout my MPhil

study, which have fostered my critical thinking and helped me to overcome many

technical problems in my study. In addition, I would like to acknowledge the

stimulating discussions and constructive suggestions by my co-supervisors: Dr Lu

Zhao and Dr Thomas Baumgartl. In particular, I have benefited a lot for the ICP-OES

training by Dr Lu’s strong analytical chemistry background. At the same time, I would

like to thank Ms Xiaohong Yang from Forensic and Scientific Services Queensland

Health for the ICP-MS and HPLC-ICP-MS analysis. I also would like to express my

gratitude to the UQ Glasshouse Manager Ken Hayes for his professional knowledge

and kind help during my glasshouse trial.

Last but not least, I would like to give my sincere thanks to the research staff and

postgraduate students in the CMLR, particularly Miss Zulaa Dorjsuren for her help in

the glasshouse experiments and Dr Xiaofang Li for his guidance in laboratory

analysis. I would also like to extend my gratitude to my friends and fellow

postgraduate students, including Yumei Du, Qi Shao, Jiajia Zheng, Fangyou, Mingrui

Yuan and Shasha Jiang for their support and friendship.

Keywords

Arsenic, magnetite, sugarcane, biochar, plants, tailings, pore water

Australian and New Zealand Standard Research Classifications (ANZSRC)

ANZSRC code: 030901, Environmental Chemistry, 30%

ANZSRC code: 050304, Soil Chemistry, 30%

ANZSRC code: 050207, Environmental Rehabilitation, 40%

Fields of Research (FoR) Classification

FoR code: 0607, Plant Biology, 30%

FoR code: 0502, Environmental Science and Management, 40%

FoR code: 0301, Analytical Chemistry, 30%

i

Table of Contents

Chapter 1 Introduction............................................................................................................ 1

1.1 Environmental significance of tailings ............................................................................ 1

1.2 Dissolution and solubility of As minerals ........................................................................ 2

1.3 Arsenic uptake by plants .................................................................................................. 4

1.4 Magnetite in copper tailings and effects on As forms and availability............................ 4

Chapter 2 Literature review and research objectives .......................................................... 7

2.1 Arsenic pollution in the natural environment .................................................................. 7

2.2 Arsenic dissolution, adsorption and transformation in the continuum of soil-solution

phase ...................................................................................................................................... 9

2.2.1 Arsenic and Fe minerals in mine tailings .................................................................. 9

2.2.2 Arsenic mineral dissolution and transformation ..................................................... 11

2.2.3 Arsenic chemical forms in tailings and factors influencing its transformation ...... 13

2.3 Arsenic in pore water: adsorption and speciation .......................................................... 15

2.3.1 Basic chemistry and chemical forms in aqueous phase and plant uptake ............... 15

2.3.2 Arsenic adsorption-desorption process in the solid-solution interface ................... 15

2.3.3 Arsenic speciation in solution regulated by microbial processes and redox

conditions ......................................................................................................................... 18

2.4 Arsenic uptake, transport and distribution in plants ...................................................... 20

2.4.1 Uptake mechanisms in roots influencing factors .................................................... 20

2.4.2 Arsenic transport and distribution in plants ............................................................ 23

2.4.3 Species diversity in As uptake and accumulation ................................................... 24

2.5 Impacts of tailings amendment on As availability and plant uptake ............................. 26

2.5.1 Organic amendment ................................................................................................ 26

2.5.2 Inorganic amendment impacts ................................................................................ 27

2.6 Organic amendment and phytostabilization of Cu mine tailings ................................... 28

ii

Chapter 3 Altered arsenic distribution in copper tailings of contrasting magnetite

content and under organic matter amendment................................................................... 30

3.0 Introduction .................................................................................................................... 30

3.1 Materials and Methods ................................................................................................... 34

3.1.1 Physicochemical analysis: pH, EC, particle size and total element concentrations 34

3.1.2 Fe Mn Al extraction ................................................................................................ 35

3.1.3 Arsenate adsorption ................................................................................................ 35

3.1.4 Arsenic fractionation ............................................................................................... 36

3.1.5 Data analysis ........................................................................................................... 37

3.2 Results ............................................................................................................................ 37

3.2.1 Physicochemical properties .................................................................................... 37

3.2.2 Distribution of Fe/Al/Mn oxyhydroxides ............................................................... 39

3.2.3 Arsenate adsorption by EHM tailings ..................................................................... 41

3.2.4 Arsenic fractionation ............................................................................................... 42

3.3 Discussion ...................................................................................................................... 47

3.3.1 Relationship between property changes induced by magnetite recovery and As (V)

adsorption in the tailings .................................................................................................. 47

3.3.2 Arsenic distribution and re-distribution in the tailings ........................................... 49

3.4 Summary ........................................................................................................................ 50

Chapter 4 Arsenic dissolution and speciation in pore water of high and low magnetite

tailings amended with organic matter.................................................................................. 52

4.0 Introduction .................................................................................................................... 52

4.1 Materials and Methods ................................................................................................... 54

4.1.1 Plant culture and treatment ..................................................................................... 54

4.1.2 Pore water sampling and chemical analysis............................................................ 56

4.1.3 Plant harvest and analysis ....................................................................................... 57

4.1.4 Data analysis ........................................................................................................... 57

4.2 Results ............................................................................................................................ 58

iii

4.2.1 Pore water properties in EHM tailings.................................................................... 58

4.2.2 Plants response to tailings ....................................................................................... 66

4.3 Discussion ...................................................................................................................... 70

4.3.1 Arsenic and Fe dissolution in pore water of LM and HM tailings amended with

organic matter .................................................................................................................. 70

4.3.2 Arsenic speciation in the pore water and Fe mineral forms .................................... 72

4.3.3 Arsenic uptake by native grass and implication for remediation ............................ 73

4.4 Summary ........................................................................................................................ 75

Chapter 5 General Discussion............................................................................................... 76

5.1 Major differences in As fractionation between LM and HM Cu tailings. ..................... 77

5.2 Mechanisms of As dissolution and speciation in pore water and plants uptake ............ 78

5.3 Conclusions .................................................................................................................... 79

Bibliography ........................................................................................................................ 81

Chapter 6 Supplementary Figures ....................................................................................... 98

iv

List of Figures

Figure 1-1: A conceptual diagram illustrating the process of Arsenic dissolution from the Fe

minerals ...................................................................................................................................... 3

Figure 3-1: The process of magnetite removal in EHM tailings and flotation test (Davey KJ

2008) ........................................................................................................................................ 32

Figure 3-2: Particle size distribution in EHM tailings (clay: 0-2 µm; fine silt: 2-6.3 µm;

medium silt: 6.3-20 µm; coarse silt: 20-63 µm; sand: >63 µm) .............................................. 38

Figure 3-3: Arsenate adsorption isotherms in EHM tailings which were fitted with the

Langmuir model ....................................................................................................................... 42

Figure 4-1: Changes of pH conditions in the pore water of the amended LM and HM tailings

during the 4 weeks of glasshouse incubation ........................................................................... 59

Figure 4-2: Pore water EC in the LM and HM tailings during the period of glasshouse

experiment................................................................................................................................ 59

Figure 4-3: Redox potentials in the amended tailings under well watered conditions in the

glasshouse experiment, which were measured in freshly collected leachate of 1 pot twice a

week after commencing treatment ........................................................................................... 60

Figure 4-4: The distribution of soluble As among different As species in the pore water of the

LM and HM tailings amended with combinations of sugarcane SR and BC, (LM refers to low

magnetite tailings, HM refers to high magnetite tailings) ....................................................... 65

Figure 4-5: Arsenic concentrations in the shoot and root of Red Flinders grass harvested in

the 4th

week after commencing treatments of organic matter amendments in the LM and HM

tailings ...................................................................................................................................... 68

Figure 4-6: The relationship between total soluble As concentration and soluble Fe

concentration in the organic matter amended LM and HM tailings from the glasshouse

experiment................................................................................................................................ 72

Figure 4-7: Iron concentration in the root of Red Flinders grass harvested at the 4th

week after

commencing treatments in the organic matter amended LM and HM tailings ........................ 74

Figure 5-1: A conceptual diagram illustrating possible mechanisms of As and Fe mineral

dissolution, transformation, adsorption and speciation and plants uptake in Cu tailings under

organic matter amendment and revegetation. Three mechanisms have been proposed (1), (2)

and (3) in the diagram which has been interpreted in the discussion. ..................................... 80

Figure 6-1: The setup of twin-pot system and pore-water sampler ......................................... 98

v

Figure 6-2: Plant growth in the amended tailings by using the twin-pot system. Plants were

irrigated by bottom-fed water via capillary suction ................................................................. 99

List of Tables

Table 2-1: Comparative risks of arsenic in microorganism, plants, animals and human .......... 8

Table 2-2: Common examples of Fe-bearing As secondary minerals ..................................... 11

Table 2-3: Arsenic levels in soils, tailings and groundwater. The data were extracted from the

literature ................................................................................................................................... 16

Table 2-4: Effect of Eh-pH conditions on As speciation in the aqueous environment (Schnoor

1996) ........................................................................................................................................ 19

Table 2-5: Influences of phosphate on arsenate uptake in different species. The experimental

information has been extracted from the literature .................................................................. 21

Table 2-6: Organic arsenic accumulated in plant tissues. ........................................................ 22

Table 2-7: Arsenic distribution in plant tissues of different plant species ............................... 24

Table 2-8: Comparison of arsenic accumulation among different plant species grown in As-

contaminated soils .................................................................................................................... 25

Table 2-9: Examples of arsenic transport from roots to shoots of different plant species....... 26

Table 3-1: Arsenic fractionation method by Wenzel ............................................................... 37

Table 3-2: Background physicochemical properties of EHM tailings SR and BC (air-dry

weight) used in the experiment, including pH water, EC, total element concentrations and

crystalline Fe minerals composition (Quantitative XRD) ....................................................... 39

Table 3-3: The distribution of extractable Fe/Mn/Al oxyhydroxides in the LM and HM

tailings, in response to organic matter treatments under direct revegetation with Red Flinders

grass ......................................................................................................................................... 41

Table 3-4: Arsenic distribution among the chemical forms in LM and HM tailings (organic

matter amended and unamended tailings). The amended tailings were incubated in a well-

watered status under glasshouse conditions for four weeks, in which Red Flinders grass was

grown ....................................................................................................................................... 44

Table 3-5: ANOVA summary of main effects and interactions of the treatment factors

(magnetite content and SR) on the distribution of As in different chemical forms in LM and

HM tailings (amended and unamended tailings) ..................................................................... 45

vi


(magnetite content and BC) on the distribution of As in different chemical forms in LM and

HM tailings (only amended tailings) ....................................................................................... 45

Table 3-7: Correlation between As fractions in the tailings and tailings properties. The data

from the amended and unamended LM and HM tailings were pooled together in correlation

analysis ..................................................................................................................................... 46

Table 3-8: Correlation among various chemical forms of As in the tailings ........................... 47

Table 4-1: The nutrient solution used to irrigate plants in the glasshouse “twin-pot” system 55

Table 4-2: Total As concentration (µg L-1

) in pore water of the LM and HM tailings amended

with organic matter .................................................................................................................. 61

Table 4-3: Total Fe concentration (mg L-1


with organic matter .................................................................................................................. 62


(magnetite content and BC) on total As and Fe concentrations in pore water of the LM and

HM tailings .............................................................................................................................. 62

Table 4-5: Concentrations of As species in the pore water collected in the 4th

week from the

LM and HM tailings amended with organic matter ................................................................. 64


(magnetite content and BC) on the distribution of soluble As among different As species in

the LM and HM tailings amended with organic matter ........................................................... 64

Table 4-7: The biomass of Red Flinders grass grown in organic matter amended LM and HM

tailings for 4 weeks .................................................................................................................. 66


(magnetite content and BC) on the plant biomass of Red Flinders grass grown in the amended

LM and HM tailings................................................................................................................. 67


(magnetite content and BC) on As concentrations in the shoot and root of Red Flinders grass

grown in the LM and HM tailings amended with biochar (BC). Sugarcane residue was used

as a basal amendment across the treatments ............................................................................ 69

Table 4-10: Correlation between As concentrations in Red Flinders grass and concentrations

of total As and As species in the pore water of the LM and HM tailings amended with organic

matter ....................................................................................................................................... 69

vii

List of Abbreviations

EHM Ernest Henry Mine

EC Electrical Conductivity

OM Organic Matter

SR Sugarcane Residue

BC Pine Biochar

ICP-OES Inductively Coupled Plasma Optical

Emission Spectrometry

HFO Hydrous Ferric Oxide

XRD X-ray Diffraction

MTL Maximum Tolerable Level

mg L-1 Milligrams per Litre

µg L-1 Micrograms per Litre

AsB Arsenobetaine

DMA Dimethylarsinic acid

MMA Monomethylarsonic acid

TMAO Trimethylarsine oxide

LSD Least Significant Difference

1

Chapter 1 Introduction

1.1 Environmental significance of tailings

Compared to other metallic mining and processing activities, base metal mines generate the

largest amount of wastes due to the total volume of ores mined and processed and the total

volume of metals produced each year (Mason et al. 2011; Mudd 2010). Base metal mine

tailings and residues contain abundant levels of heavy metals, metalloids, radionuclides and

other pollutants, which in recent years, have entered natural environments at a much greater

pace and in more geographical locations, for example, Australia, Africa, China and South

America (Gordon 2002; Li 2006; Power et al. 2011; Rogich and Matos 2008). This trend is

worsening due to the largely low grades of mineral ores and the rapidly increasing volume of

metal extraction (Crowson 2012; Mudd 2007; Power et al. 2011; Prior et al. 2012). The

tailings without proper management pose great environmental risks and affect environmental

quality and human health (Dudka and Adriano 1997; Mendez and Maier 2008).

Levels of arsenic (As) are present in abundance in copper, gold and uranium tailings (Dold

and Fontboté 2001; King et al. 2008). There are many As-bearing minerals in Cu tailings,

such as arsenopyrite (FeAsS), chalcopyrite (CuFeS2), enargite (Cu3AsS4) and tennantite

(Cu12As4S13) (Dold and Fontboté 2001; Filippou et al. 2007; Mielczarski et al. 1996).

Exposure to arsenic from tailings threatens wildlife and human health, especially children

who ingest the polluted soil accidentally (Rodriguez et al. 1999). Without proper

management, arsenic in tailings can be easily leached into the groundwater and drinking

water. For example, As in groundwater has posed a great threat to people in Bangladesh who

have to drink the As-polluted water unaware of the risks (Chowdhury et al. 2000).

As a result, tailings rehabilitation is necessary to diminish pollution risks and prevent the off-

site transport of tailings particulates and seepage water. Land polluted with heavy metals and

metalloids may be rehabilitated by physical, chemical and biological technologies (Khan et al.

2000). Methods such as vitrification, land filling, chemical treatment and electro kinetics

have recently been applied to the contaminated land (Robinson et al. 2003). However, these

methods have some serious disadvantages, including short-term effectiveness, unsuitability

for large areas/volumes of wastes or contaminated land, high implementation and post-

remediation maintenance costs and residual chemical risks following treatment.

Phytoremediation, which involves using plants to remove, degrade and immobilise

2

contaminants, has been advocated as the most effective way to rehabilitate mine tailings and

heavily contaminated land due to its low-cost and long-term sustainability (Hughes et al.

1996; Padmavathiamma and Li 2007). Phytostabolization using a native plant ecosystem is a

common requirement in mine plans and demands proper root zone reconstruction and

rehabilitation of the root zone functions. Native plants that are well adapted to the local

environment can be cultivated in the contaminated mine site. However, before effective

phytotabilization can be developed, it is necessary to understand the hydro-geochemistry of

the tailings and remediation effectiveness in relation to plant responses.

1.2 Dissolution and solubility of As minerals

Although As contained in many primary and secondary minerals may not be directly

bioavailable to plants and animals, biogeochemical processes catalysed by chemical and

microbial factors can transform insoluble As forms in the minerals into readily exchangeable

forms, releasing anionic forms of As into pore water in soil and mine wastes such as tailings

(Bauer and Blodau 2006; Harvey et al. 2002). The availability of As can be influenced by

combined effects of many environmental factors including pH, rainfall, redox conditions (Eh),

and geological factors including mineral composition and soil types. In general, As solubility

in soil increases with rising pH from acidic to neutral (pH 4-8), which is similar to the

solubility of phosphate in soil (Smedley and Kinniburgh 2002). As a result, arsenic

mobilisation can be inhibited by the addition of phosphate due to their similar chemical

characteristics and competition for adsorption sites at containing minerals including Fe

oxides and Fe-oxyhydroxides (Manning and Goldberg 1996; Smith et al. 2002). In soil, As is

more likely to be released from the Fe oxyhydroxides under anaerobic conditions (0-100 mv)

due to the associated reduction and dissolution of Fe3+

(Masscheleyn et al. 1991). Arsenate

has a high affinity to the Fe oxyhydroxides and can be adsorbed in the inner surface of these

minerals (Waychunas et al. 1993). In soil, arsenic desorption from Fe oxyhydroxides is also

mediated by the activities of Fe-oxidising and reducing bacteria which directly catalyse iron

dissolution and As release into the pore water indirectly (McCreadie et al. 2000; Mendez et al.

2008; Morin and Calas 2006).

In copper tailings under direct revegatation, arsenic dissolution is predominantly influenced

by the adsorption of iron oxyhydroxides and the reduction and dissolution of iron by

microbial activities. In the weathering process, these As-Fe primary minerals can be

transformed into secondary minerals, releasing As (H2AsO4-) and Fe (Fe

2+) into the pore

3

water. In tailings, the oxidation of primary minerals such as the arsenopyrite (FeAsS) can

release high levels of As into the water (up to 72 mg L-1

) and soil through acid mine drainage

(Williams 2001). Organic matter can supply organic carbon resources (C6H12O6), which can

be utilised by the Fe oxidising or reducing bacteria as electron acceptors (Lovley et al. 1998;

Lovley and Phillips 1988). This stimulates the iron dissolution from the minerals, thereby

releasing As into the pore water, which is bonded in the Fe oxyhydroxides.

In pore water, As transformation among inorganic (including arsenate (AsO43-

), arsenite

(AsO33-

)) and organic forms (including monomethylarsonic acid [MMA] (CH3AsO(OH)2 ),

dimethylarsinic acid [DMA] ( (CH3)2AsO(OH)), trimethylarsine oxide [TMAO] ( (CH3)3AsO)

and penyl arsenic (C6H5AsO(OH)2) is controlled by redox conditions and processes mediated

by relevant reducing/oxidising bacteria. Arsenate can be transformed to arsenite under

reducing conditions while arsenite can be oxidised into arsenate under oxidising conditions

(Zhao et al. 2010). Inorganic As forms can be converted by microorganisms into the organic

As forms (Cullen and Reimer 1989; Zhang and Selim 2008).

As a result, As chemical forms and solubility in mine tailings may be altered by various

amendment strategies and revegetated plant species. In particular, the addition of organic

matter may mediate bacterial activities and transform Fe and As minerals in the tailings.

Moreover, the presence and content of Fe-minerals and their transformation may significantly

alter As-adsorbing capacity and thus the distribution of As in the pore water of amended

tailings. It is necessary to investigate the influence of both altered mineralogy (e.g. removal

of magnetite) and organic matter amendments in copper tailings on As chemical forms and

Fe-oxidizing &

reducing

microorganisms

As-Fe (III)

primary

minerals

Carbon resources

(C6H12O6)

Increase

Organic matter

Supply

Catalyse Release As (V) (H2AsO4-)

Ferrous (Fe2+

)

Figure 1-1: A conceptual diagram illustrating the process of Arsenic dissolution from the Fe

minerals

4

plant uptake, in order to assess potential risks of As mobilisation in the pore water during the

process of revegetation.

1.3 Arsenic uptake by plants

Arsenic is a nonessential element to plants. Plants roots can absorb both inorganic and

organic As species. The uptake of arsenate and phosphate shares the same pathway (Asher

and Reay 1979; Ullricheberius et al. 1989). In contrast, arsenite can be taken up by roots in

the form of the neutral molecule (H3AsO3) under reducing conditions such as those in paddy

soils (Inskeep 2002; Xu et al. 2008). Organic As species such as MMA and DMA are also

present in soils and water, but they cannot be as effectively absorbed by roots compared to

the inorganic arsenate and arsenite (Carbonell-Barrachina et al. 1998; Marin et al. 1992; Raab

et al. 2007). Current research findings have confirmed that arsenite is the predominate specie

in plant tissues (Pickering et al. 2000; Raab et al. 2005; Xu et al. 2007). When As is

transported from roots into shoots, the majority of As species in the xylem are reduced into

arsenite (Zhu and Rosen 2009).

Many studies have investigated As uptake by crops including rice, wheat and maize in

contaminated soils and water (Bai et al. 2008; Meharg 2004; Williams et al. 2007). Different

species tend to have different As tolerance and associated physiological mechanisms. For

example, Gulz (2005) found that ryegrass grown in a contaminated soil accumulated 255 mg

kg-1

As in the root but merely 11 mg kg-1

As in the shoot. In contrast, the hyper accumulator

plant-Pteris vittata (brake fern) has a high As transfer factor from root to shoot, resulting in

as much as 23000 mg As kg-1

in the shoot (Ma et al. 2001). In general, As is more likely to

remain in roots due to the presence of As-adsorbing Fe-plaque at the root surface under

oxidising conditions (Zhao et al. 2010). In comparison with crop/pasture species, As uptake

by native plant species grown in amended mine tailings have received relatively little

research attention.

1.4 Magnetite in copper tailings and effects on As forms and availability

Large amounts of tailings have been accumulated in copper mines, including Ernest Henry

Mine (EHM) despite being a relatively new mine (Siliezar J 2011). Arsenic bearing minerals

are commonly associated with Cu-ore minerals and gangue materials, thus resulting in large

amounts of As in the tailings (Drahota and Filippi 2009). For example, Cu (chalcopyrite)-

gold (Au) mineralisation at EHM occurs mainly within the magnetite-biotite-calcite ± pyrite

matrix of a pipe-like breccia body (Siliezar J 2011). The EHM ore contains high contents of

5

magnetite, which could have economic value under favourable market conditions. As a result,

one option recommended to EHM is to recover and stockpile magnetite from the tailings by

altering the ore processing circuit and reprocessing the tailings after Cu-flotation.

Fe-minerals (e.g., iron oxides and oxyhydroxides) play an important role in As adsorption

(Koo et al. 2012; Waychunas et al. 1993; Wenzel et al. 2002). As a result, magnetite removal

may increase As mobilisation due to the perceived reduction of the As adsorption capacity.

The distribution of As forms and As bio-availability in Cu tailings may be altered by the

removal of iron oxides (such as magnetite) to favour its distribution into the pore water

(Kundu and Gupta 2006). When the tailings are subject to organic amendment and

revegetation, As minerals in Cu-tailings may be transformed into chemical forms under

microbial and chemical conditions in the rhizosphere, thus posing much greater ecological

risks. As a result, it is necessary to investigate whether organic matter (OM) amendment

could stimulate more As dissolution in the low magnetite tailings than in the high magnetite

tailings (i.e., before magnetite recovery).

The expected changes in As forms and bioavailability form the central theme of the present

research project. Experimental investigations were conducted using Cu-tailings collected

Ernest Henry Mine, located in Cloncurry, North Queensland, Australia.

In the present rehabilitation plan, the EHM tailings storage facility is to be rehabilitated into

native pasture land. Thus grazing wild animals and uncontrolled farm stock (e.g, goats and

cattle) may be exposed to the As accumulated in herbage, leading to As intake in their diets.

In particular, the altered mineral processing in EHM to remove the As-sink mineral magnetite

may enhance the availability of As for plant uptake. However, it is still unclear whether

magnetite removal and organic matter amendment could increase As distribution into the

pore water and plant uptake by native grass species grown in copper tailings. Hence, the

study will focus on the effects of magnetite removal on As forms in copper tailings and As

uptake and accumulation by Australian native grasses.

The detailed aims are to investigate the amended and revegetated Cu tailings with contrasting

magnetite contents:

1. As adsorption characteristics in the tailings and As distribution among different

chemical forms and in relation to the tailings magnetite contents and other chemical

properties;

6

2. As speciation in the pore water, in relation to organic matter amendment in the

tailings;

3. As accumulation and distribution in the native grass grown in two tailings under the

direct revegetation.

The two types of tailings were firstly characterised for physical and chemical properties,

including particle size distribution, pH, EC, Eh, As fractionation, As (V) adsorption and Fe

forms (Chapter 3). In order to understand As bioavailability in the tailings under direct

revegetation, chemical forms of As in the solution phase were analysed by speciation analysis

(Chapter 4). The As speciation results were interpreted in relation to As accumulation and

distribution in the tissues of native grass (Red Flinders Grass) (Chapter 4). On the basis of the

present research findings, the mechanisms of As mobilisation in copper tailings and its uptake

by plants are discussed briefly together within the context of the literature review (Chapter 5).

7

Chapter 2 Literature review and research objectives

2.1 Arsenic pollution in the natural environment

Arsenic is a metalloid and is naturally present in the earth’s crust with an average

concentration of approximately 5 mg As kg-1

(Semeraro et al. 2012). Natural As in the

environment may come from sources such as weathering of rocks and volcanic emissions,

however elevated levels of As in soil and water come from anthropogenic sources, for

example mining, agricultural processing, forestry and urban wastes (Fitz and Wenzel 2002;

Matschullat 2000). Generally, natural soils normally contain < 10 mg As kg-1

while As

contaminated soils may contain up to more than 1000 mg As kg-1

(Bañuelos and Ajwa 1999;

Cancer 2004). In Australia, more than 10,000 soil sites have been identified as As polluted

(Moreno-Jimenez et al. 2012). According to the NEPC guidelines (NEPC 1999), health

investigation limit of total As concentration in residential garden soils is 100 As mg kg-1

in

soil and 200 mg As kg-1

in parks and recreational open spaces, while the ecological

investigation level in urban soil is as high as 200 mg As kg-1

soil. In recent years, large

volumes of base metal mine tailings rich in As-minerals have become significant sources of

As pollution in natural environments as the tailings contain many As-bearing primary and

secondary minerals (Drahota and Filippi 2009; Mudd 2010). Arsenic pollution in the

environment can cause a range of adverse impacts on the quality of soil, water and plants

(food), leading to human health problems through the intake of food and water (Table 2-1).

People who are exposed to As-contaminated water and food may be at risk of As poisoning,

resulting in diseases of the skin, digestive and nervous systems (Kapaj et al. 2006). As a

result, effective management solutions to minimise and prevent the pollution impacts of mine

tailings are critical. Phytostabilization has been advocated as a cost-effective and sustainable

management option for tailings management as part of the mine closure plan (Huang et al.

2012).

8

Table 2-1: Comparative risks of arsenic in microorganism, plants, animals and human

Catalogue Species or

tissues

As levels Symptom Reference

Microorganisms Escherichia coli 2 mmol L-1

sodium arsenite

Intracellular

ROS, LPO and

DNA damage

(De et al. 2012)

Plant Rice (Oryza

sativa L.)

10-70 mg kg-1

in

soil and 0.13 mg

kg-1

in irrigating

water

Yield declined

severely from 7-

9 to 2-3 t ha-1

(Panaullah et al.

2009)

Ladder brake

(Pteris vittata L.

Hyper-

accumulator)

500 mg kg-1

Biomass

reduced by 64%

(Tu and Ma

2002)

Soil fauna Earthworm

(L.rubellus)

2000 mg

arsenate kg−1

Died in 28 days (Langdon et al.

1999)

Animal Male BALB/c

mice

3.2 mg L-1

As

drinking water

Significant liver

disease in 9

months

(Santra et al.

2000)

Human People in West

Bengal, India

Arsenic in

drinking water

50-1188 µg L-1

Skin lesion

disease

(Ghosh et al.

2007)

Further to the air-borne pathways, the likelihood of As in base metal mine tailings entering

the natural soil and water environments depends on As solubility in the pore water and

seepage water. This involves the interwoven processes of dissolution and transformation of

As-bearing minerals and As speciation in the solid-solution interfaces regulated by physical,

chemical and biological factors (Morin and Calas 2006; O'Day 2006; Rosso and Vaughan

2006). Inorganic As in the solution phase may be re-adsorbed by secondary Al/Fe/Mn

minerals (such as oxides and oxyhydroxides) to form As-secondary minerals of low solubility,

9

but which may be released into the pore water under suitable biogeochemical conditions

(Drahota and Filippi 2009; Morin and Calas 2006). As a result, it is important to characterise

As forms, phytoavailabilty and biogeochemical processes involved in As transformation and

speciation in tailings in order to understand the potential risks of As distribution and

mobilisation in Cu-tailings subject to amendments and revegetation.

2.2 Arsenic dissolution, adsorption and transformation in the continuum of soil-solution

phase

Arsenic in primary and secondary minerals in mine wastes (i.e. the solid phase) may be

released into soil solution or pore water through dissolution processes driven by

biogeochemical factors (e.g. acidification, oxidation and reduction). The soluble As may be

re-adsorbed by Fe-oxides and oxyhydroxides and other minerals and transformed into less

soluble minerals (Al-Abed et al. 2007; Bhattacharya et al. 2007; Drahota and Filippi 2009;

Duker et al. 2005). Once in aqueous solution, As species can be present in both inorganic and

organic forms, depending on the chemical and microbiological conditions determining the

distribution of As among different chemical species (i.e. chemical speciation). Arsenic

speciation in the solution phase determines plant availability and bio-toxicity in the tailings or

soil environment (Stoltz and Greger 2002; Visoottiviseth et al. 2002; Xie and Huang 1998).

To understand possible risks of As toxicity in amended mine tailings under revegetation, it is

useful to understand the kind of As minerals that are likely to be present.

2.2.1 Arsenic and Fe minerals in mine tailings

So far more than 300 kinds of arsenic minerals have been discovered in several geological

deposits, including volcanogenic sulphide, epithermal gold and porphyry copper ores

(Arehart et al. 1993; Çiftçi et al. 2005; Leybourne and Cameron 2008). In addition to other

mineral substrates including arsenite and metal mixture deposits, arsenates account for

approximately 60%, sulphides and sulphosalts 20%, oxides 10%, while the other mineral

substrates are arsenite and metal mixtures (Parshley 2001). In nature, arsenic is

predominantly found in minerals associated with As anions such as As2 and AsS (Ravenscroft

et al. 2011). Arsenic anions, which are associated with heavy metals such as Fe (arsenopyrite),

Co (cobaltite), Cu (chalcopyrite) and Ni (gersdorffite), are the main primary arsenic-bonding

minerals (Drahota and Filippi 2009).

Ernest Henry Mine is located in 38 km north-east of the Mt Isa-Cloncurry mineral district of

North-West Queensland, with an estimated annual copper production of 100,000 tonnes,

10

which has generated about 10,700,000 tonnes of tailings in the environment (Siliezar J 2011).

EHM copper tailings contain approximately 21% Fe, 38.2% SiO2, 8.5% Al2O3, 5.3% K2O,

2.3% S, 0.17% P, 0.077% Cu and 0.033% As (Siliezar J 2011). On the basis of the production

rate of tailings, there are approximately 350 tonnes of As already being released into the

environment each year, although the tailings have been mostly contained in a purposely

designed tailings storage facility (TSF). Magnetite is the dominant iron oxide species in the

chalcopyrite ore, which accounts for 20% to 25% in the tailings. The economic potential of

magnetite triggered the re-processing of copper tailings to recover magnetite in December

2009 (Siliezar J 2011). However, this change in mineral processing may alter the mineral

composition in the resultant tailings, particularly the As-adsorbing Fe-minerals such as

magnetite and its derivatives, and the quantity and chemical forms of As, leading to different

risk potentials in the environment.

As not all As primary minerals are stable, they can be converted to secondary As minerals in

the progress of weathering under suitable biogeochemical conditions (Foster et al. 1998). The

weathering of As-bearing minerals progressively occurs at the surface of the minerals

exposed to the atmospheres and water, resulting in the formation of oxides and other anions

(Waychunas et al. 1993). A wide range of secondary As associated minerals have been

detected in contaminated soils and mine tailings, including As oxides, Fe arsenates, Fe

sulphides, Ca, Mg arsenates and other metal arsenates (Filippi et al. 2009; Paktunc et al. 2004)

(Table 2-2). Recently, As bonding with these secondary minerals has been further clarified

using modern analytical tools including micro-Raman Spectroscopy and X-ray Absorption

Spectroscopy techniques. These studies revealed the importance of Fe-minerals in As

immobilisation (Courtin-Nomade et al. 2002; Zänker et al. 2002; Zanker et al. 2003). Some

of the mineral-bound As (via the dissolution process) may be mobilised into pore water in

tailings/soil, which can then be absorbed by plant roots and/or transported offsite via

seepage/runoff. Because As mobilisation and adsorption are closely coupled with the content

and properties of Fe-minerals, the following sections will particularly address the processes

involved in the dissolution and transformation processes of both As and Fe minerals in the

tailings. The As mobilisation and adsorption may be simultaneously influenced by in situ bio-

geochemical conditions in tailings under amendment and phytostabilization.

11

Table 2-2: Common examples of Fe-bearing As secondary minerals

Secondary

As minerals

Formula Mine sample References

Scorodite FeAsO4·2H2O California mine

tailings

(Foster et al. 1998;

Paktunc et al.

2004)

Alunite KAl3(SO4,AsO4)2(OH)6 Mother Lode Gold

District of California

(Savage et al.

2000)

Jarosite KFe3(SO4,AsO4)2(OH)6 Mother Lode Gold

and Ketza River

Mine tailings

(Paktunc and

Dutrizac 2003;

Savage et al.

2005)

Tooeleite Fe(III)6(AsO3)4(SO4)(OH)4.4H2O Carnoulès mine,

Gard, France

(Morin et al. 2003)

Pharmaco-

siderite

K2Fe4(AsO4)3(SO4)(OH)4.4H2O Echassieres, France (Morin et al. 2002)

Arsenioside

rite

Ca2Fe3(AsO4)3.3H2O Ketza River Mine

tailings

(Paktunc et al.

2004; Paktunc et

al. 2003)

2.2.2 Arsenic mineral dissolution and transformation

Dissolution of As and Fe minerals

Arsenopyrite (FeAsS) is one of the most common primary As-bearing minerals, which can be

readily oxidised to release arsenic ions to pore water (Drahota and Filippi 2009).

Arsenopyrite oxidisation occurs upon being exposed to oxygen and water, resulting in the

generation of sulphuric acid and associated dissolution of Fe and As and the subsequent

formation of iron oxides and hydroxides and arsenic anions (Eq. 1) (Shuvaeva et al. 2000).

The oxidation of pyrite minerals can also be catalysed indirectly by bacteria including

Acidithiobacillus ferrooxidans (Morin and Calas 2006).

12

FeAsS + 3.5O2 + 4H2O → Fe(OH)3 (s) + HAsO42-

+ SO42-

+ 4H+ (1)

The dissolution and transformation of AsFe sulphides are complex and affected by prevailing

biogeochemical conditions, such as pH, redox potential and microbial activities. Under acid

(pH <3) conditions, arsenopyrite can be transformed into scorodite (FeAsO4.2H2O) which is

the most common As secondary mineral (Drahota and Filippi 2009). Scorodite is present in

naturally weathered rocks (Utsunomiya et al. 2003), contaminated soils (Morin et al. 2003;

Pfeifer et al. 2004) and many mine tailings (Craw et al. 2002; Davis et al. 1996; Mahoney et

al. 2005). The dissolution of scorodite occurs when the pH is less than 3 (Frau and Ardau

2004; Moldovan and Hendry 2005), releasing As5+

into the aqueous phase. This process is

described in the following reaction (Dove and Rimstidt 1985).

FeAsO4·2H2O(s) + H+

(aq) → H2AsO4−

(aq) + Fe (OH) 2+

(aq) + H2O (l) (2)

Under reducing conditions in soil or tailings, sulphate-reducing bacteria can catalyse the

reduction of As-hematite into ferrous iron and release arsenious acid (Eq. 3) (McCreadie et al.

2000).

2 Fe2 O 3 .H3AsO3+CH2O+7H

+→ 4Fe

2++HCO3

-+4H2O+2H3AsO3 (3)

Biotic and abiotic factors affecting mineral dissolution and As adsorption process

The soluble As pool in pore water of the tailings results from the dissolution of As-bearing

minerals and the adsorption of soluble As by positively charged mineral surfaces, which are

closely influenced by complex interactions of many biogeochemical factors. Both abiotic

(such as pH and redox) and biotic (such as oxidising/reducing bacteria) factors play

significant roles in the adsorption and dissolution of As/Fe minerals in soil and tailings. Soil

or tailings pH conditions can strongly influence the As dissolution from the Fe-oxides and

Fe-oxyhydroxides, although the exact relationship between pH change and As adsorption by

Fe-minerals may vary with mineral types for a given pH range (Al-Abed et al. 2007; Grossl

and Sparks 1995; Manning and Goldberg 1997). Manning and Goldberg (1997) found that

arsenate adsorption on goethite, magnetite and hematite increased with rising pH in the range

of strongly acidic to slightly alkaline (pH 3-8). Arsenate sorption on goethite decreased from

slightly acidic to alkaline pH (6-11) (Grossl and Sparks 1995). In a separate study, both

arsenate and arsenite sorption on goethite decreased with increasing pH from neutral to

strongly alkaline (i.e. pH 7-14) (Matis et al. 1997). In contrast to goethite arsenite sorption by

13

magnetite increased with increasing pH from acidic to moderately alkaline (pH 9), but it

decreased in more alkaline conditions at pH >9 (Dixit and Hering 2003). From these studies,

we can conclude that the adsorption of As by Fe-minerals increases with increasing pH from

strongly acidic to neutral, but decreases with a further pH increase into alkaline and strongly

alkaline conditions. This is influenced by the characteristics of Fe minerals and As species

(i.e. As (V) vs As (III)).

Redox conditions in the tailings or soil may alter the composition of secondary minerals that

have an affinity for As anions. In most cases, As solubility increases with the dissolution of

iron oxyhydroxides in soil under reducing conditions (e.g. Eh 0-100 mv) (Masscheleyn et al.

1991). Other As forms including HAsS2, arsine, and arsenic metal may appear under the

extreme reducing and low pH conditions (Mudhoo et al. 2011), but these species are not

stable and thus negligibly relevant to plant uptake in aerobic conditions (Ning 2002).

Some Fe-oxidizing bacteria in mine tailings can mediate As transformation and adsorption

(Liu et al. 2013). For example, Gallionellaferuginea and Leptothrixochracea can oxidize Fe2+

to Fe3+

(Katsoyiannis and Zouboulis 2004; Zouboulis and Katsoyiannis 2005). Lovely (1998)

found that the reduction of Fe (III) oxides could be stimulated by the Fe-reducing bacteria

(Geobater metallireducens) by adding humic substances that act as terminal electron

acceptors. In this process, As-Fe minerals can be dissolved to release Fe2+

and associated As

anions from the surface of mineral particles into the solution phase, if it does not coincide

with the adsorption by other minerals.

In summary, the bio-geochemical conditions in the tailings under revegetation can be altered

with the amendment of organic matter, plant roots which host rhizosphere bacteria and water

availability. This is in addition to mineralogical changes resulting from ore/tailings

processing. Consequently, the dissolution of As-bearing minerals and As adsorption by

secondary Fe minerals (including magnetite and its derivatives after complex redox reactions)

may be altered by the removal of Fe-minerals from tailings and the addition organic matters

in the EHM tailings under revegetation.

2.2.3 Arsenic chemical forms in tailings and factors influencing its transformation

Total As concentration in the solid phase is not the appropriate standard to evaluate As risk to

the environment or human health, as not all As forms in soils or tailings are bio-accessible

(Koch et al. 2007). Based on the extractability of different chemical reagents, As risks in the

14

environment are related to its chemical forms in the solid phase. Arsenic chemical forms are

commonly categorised into: soluble, phosphate exchangeable, organic matter associated and

As, Fe, Al, Mn, Ca absorbable (Kim et al. 2003; Quazi et al. 2010; Salomons and Förstner

1980; Wenzel et al. 2001). For example, As bound by Fe/Al/Mn oxidises and oxyhydroxides

are considerably more stable than the soluble-As and exchangeable-As, which are more bio-

accessible under certain pH-Eh conditions and in microbe-mediated processes (Bauer and

Blodau 2006; Masscheleyn et al. 1991; Sarkar and Datta 2004a; Sarkar and Datta 2004b). In

contrast, phosphate exchangeable-As can be released into the pore water when the levels of

phosphate anions in the pore water are elevated due to its competition with arsenate for

sorption sites on the iron oxides (Cao et al. 2003; Lenoble et al. 2005; Manning and Goldberg

1996). As a result, detailed fractionation analysis of As chemical forms has been commonly

used to characterise As distribution in the solid phase of soil and tailings to assess its

potential mobility and plant availability.

Tailings remediation is a prerequisite to phytostabilization with suitable plant species, for

example adding organic matter to improve physical and chemical conditions in the tailings is

a common practice to improve rooting conditions for plant growth. Organic matter can

provide labile carbon and enhance the microbe activity in the tailings. Organic matter and its

derivatives from decomposition can not only directly influence As chemical forms through

competing functional groups of OM (such as phenolic, carboxyls, hydroxyls, etc.), but also

catalyse the transformation of Fe-primary minerals into secondary minerals such as Fe-

oxyhydroxides via microbes-mediated processes by providing labile carbon for microbes in

the tailings (Bhattacharya et al. 2007; Redman et al. 2002). Low pH and high organic matter

conditions may lead to increased proportions of exchangeable As (Giménez et al. 2007). The

dissolution of Fe minerals in tailings can be greatly catalysed by Fe-reducing microorganisms

and the As sorbed to the Fe minerals can be released into the water due to the loss of the Fe-

bonding phase (Dong et al. 2000; Giménez et al. 2007; Yamaguchi et al. 2011). As a result,

following organic matter amendment, As could dissolve from the Fe oxide surface as Fe

dissolution occurs due to the stimulated activity of Fe-reducing bacteria. For example,

Kalbitz and Wennrich (1998) found that there was a positive correlation between water

soluble As and dissolved organic carbon in the pore water of a wetland soil. Organic matter

amendment therefore may change the distribution of As in different chemical forms in

tailings under revegetation. It is necessary to understand the influences of organic matter

amendment on As distribution in the tailings to ensure informed management practices.

15

2.3 Arsenic in pore water: adsorption and speciation

2.3.1 Basic chemistry and chemical forms in aqueous phase and plant uptake

In nature, As is present in four common chemical valences, -3, 0, +3 and +5, among which

metalloid arsines (zero valences) are unstable in oxidising conditions (Mudhoo et al. 2011).

Arsenic in different valences can be present simultaneously in the solid phase of tailings or

soil, in both organic and inorganic forms (Ramesh et al. 2007; Zobrist et al. 2000). Arsenic

(V) and As (III) are the most common inorganic forms in the natural environment, which are

inter-converted under suitable redox conditions (Tripathi et al. 2007). Arsenate (H2AsO4- and

HASO42-

) are the predominant As species in aerobic soil (oxidizing environment), but arsenite

(mainly in the form of H3AsO3 at pH<9) accounts for the majority of total As under anaerobic

conditions (i.e. reducing environment), for example in paddy soils (Fitz and Wenzel 2002;

Marin et al. 1993). In aerobic conditions, arsenite can be oxidised to arsenate, while arsenate

can also be reduced back into arsenite in a reducing environment (Zhao et al. 2010).

Inorganic As species are bioavailable for plant uptake and are more toxic to plants than

organic As forms in soil and water.

Arsenate (pKa1 = 2.3, pKa2 = 6.8, and pKa3 = 11.6) and arsenite (pKa1 = 9.2, and pKa2= 12.7)

have different dissociation behaviour (Larsen and Hansen 1992). At neutral pH conditions,

arsenate appears as the anion form of H2AsO4- and HASO4

2-, while arsenite mainly exists as

molecular H3AsO3 at pH <9.2. Low pH and high organic matter conditions may lead to

increased proportions of exchangeable As (Giménez et al. 2007). In soil solutions at pH 4 - 8,

H3AsO3, H2AsO4-, and HASO4

2- are the most stable and prevalent chemical forms or species

in the aqueous phase (Smith 2007). In slightly reducing (250 mv-400 mv) and low pH (<4)

conditions, arsenious acid or arsenite (H3AsO3) may also account for a significant proportion

in the inorganic As pool, but the proportion of H2AsO3- can increase with rising pH up to 9.2;

with further pH increases HAsO32-

appeared again in the solution (Mudhoo et al. 2011). As a

result, the pH and redox conditions closely influence the speciation of inorganic As (i.e. As

(III) and As (V)) in the solution phase, while the conversion of inorganic into organic forms

of As in solution is closely influenced by microbe-catalysed processes.

2.3.2 Arsenic adsorption-desorption process in the solid-solution interface

Arsenic dissolution into the pore water of the solid phase such as tailings and contaminated

soils is a prerequisite to its transport in the seepage and/or surface run-off water. Many

studies have reported the different levels of As in contaminated soils, tailings and ground

16

water (Table 2-3). Under aerobic conditions, total As concentration in the soil solution of

natural soil is less than 50 nM, but it can be elevated to as high as about 2 µM in pore water

of contaminated soil (Waychunas et al. 1993; Wenzel et al. 2002). In paddy soils, its

concentration ranges from 0.01 to 3 µM (Zhao et al. 2009).

Table 2-3: Arsenic levels in soils, tailings and groundwater. The data were extracted from the

literature

Sample types As (mg kg-1

or

mg L-1

)

pH Reference

Contaminated soil

Pb-Zn contaminated soil in

Greece

963 8.65 (Vaxevanidou et al.

2008)

Kidsgrove contaminated soil

in UK

59.5 7.40 (Hartley et al. 2004)

Ron Phibun contaminated

soil in south Thailand

135-510 4.6-5.1 (Francesconi et al.

2002)

Mine tailings

La parrilla tailings Southwest

Spain

995-1280 2.85-3.6 (Anawar et al. 2006)

Cu-Pb-Au tailings in Santa

Maria Mexico

107-2206 6.5-8.5 (Razo et al. 2004)

Pb-Zn tailings in Bathurst,

New Brunswick, Canada

2180 3.7 (Wang and Mulligan

2009b)

Ground water

Groundwater in Lakshmipur,

Bangladesh

1.56 7.0 (Anawar et al. 2002)

Groundwater in West Bengal,

India

0.05-3.7 6.6-7.5 (Stüben et al. 2003)

Arsenic adsorption and soluble As concentration in soil or tailings can be influenced by

chemical conditions (such as pH) and mineralogy (weathering and microbial processes).

Generally, the desorption of arsenite and arsenate from the solid phase into solution increases

with increasing pH to neutral-alkaline conditions (Beesley et al. 2010). There is a “V” shaped

relationship between As adsorption and pH condition: As adsorption is low in neutral

conditions, but high at pH >8 and <4 (Masscheleyn et al. 1991). Arsenic desorption from clay

17

soils increases with increasing pH in the range of 4-8, while the maximum arsenite desorption

exists in the pH 8~10 condition (Zhang and Selim 2008). Under high pH (8-10) conditions,

arsenite is more likely to be associated with metal oxides than arsenate, however, when pH

falls below 8, arsenate shows a higher adsorption rate in soils and minerals (Goldberg 2002).

By considering pH condition alone, As in neutral-alkaline Cu-tailings may pose a greater risk

than that in acidic tailings. This certainly adds to the justification of investigating As risks in

the neutral Cu tailings with magnetite removal and under revegetation.

However, the pH-As solubility relationship in soil and tailings is strongly influenced by the

presence of amorphous Fe-minerals, which can be formed from the reduction and oxidation

of crystalline Fe-minerals under cyclic wet-dry conditions. As a result, when investigating the

dissolution and transformation processes of As-bearing minerals in tailings, it is necessary to

consider not only the pH conditions in pore water, but also the presence of Fe-minerals in

tailings. Soils rich in Fe have a high As adsorbing capacity (Wenzel et al. 2002). Amorphous

and crystalline minerals in soils, such as oxides and oxyhydroxides of iron (Fe), manganese

(Mn), and aluminium (Al), may closely influence As speciation and mobilisation through As-

adsorption process. Among Fe-oxides, goethite exhibited a stronger As sorption capacity than

that of magnetite (Bowell 1994). Arsenate adsorbed by iron oxides/oxyhydroxides can be

released into pore water and reduced into arsenite under reducing and microbial conditions,

coupled with the reduction of Fe (III) in iron oxides/oxyhydroxides into Fe (II) (Langner and

Inskeep 2000).

In the mineral weathering process, amorphous Fe oxides/oxyhydroxides such as ferrihydrite

or ferric iron (HFO) may be transformed to crystalline forms (such as hematite, goethite and

magnetite) by bacteria, which results in the reduction of a reactive surface area (Hansel et al.

2003; Zachara et al. 2002). Therefore, the transformation of Fe oxides and oxyhydroxides

would affect the adsorption capacity of As and thus its mobility in the aqueous phase. In the

meantime, the surfaces of crystalline Fe oxides may be modified and/or precipitated with

amorphous Fe-oxyhydroxides due to the redox processes and microbe-mediated processes.

The surface coating of ferrihydrites on Fe-oxides and pyrites was observed in Cu-Pb/Zn mine

tailings under microscopic examination (Forsyth 2010). The crystalline Fe oxides may be

dissolved at the surface via microbe-mediated processes; under reducing conditions the

soluble Fe2+

can be re-precipitated as secondary Fe minerals (e.g. ferrihydrites) with

increased As adsorption sites and capacity. For example, magnetite could be dissolved by the

18

bacteria Shewanella putrefaciens and release Fe2+

, but the Fe2+

would be associated with

HCO3- and generate Fe oxides such as siderite (Dong et al. 2000; Kostka and Nealson 1995).

The Fe dissolved from the surface of magnetite particles could form precipitates of Fe oxides

and oxyhydroxides, which coat the surface of magnetite particles, and thus increase the As

adsorption capacity and decrease the distribution of As in solution phase. As a result, the

content of crystalline Fe-minerals (such as magnetite) in tailings would influence As

adsorption due to the likely transformation of the Fe-minerals into the amorphous Fe

secondary minerals when the tailings are amended with organic matter under revegetation. As

a result, the reduction of magnetite content in the Cu-tailings after reprocessing at EHM

would decrease the As-adsorption capacity in the new tailings under revegetation, thus

resulting in elevated soluble As distribution in the pore water. This hypothesis will be

investigated in a glasshouse incubation experiment within present project.

In the present project, the magnetite contents in copper tailings at EHM have been

significantly reduced from 18-31% to less than 5% (EHM information) for economic

purposes. From the reviewed information, the reduction of magnetite content in the tailings

would decrease As adsorption capacity and increase As availability for plant uptake as there

would be lower As-adsorbing sites in the solid phase from lower contents of amorphous Fe-

oxyhydroxides formed in the low magnetite tailings. As a result, we hypothesise that

decreasing magnetite in the Cu-tailings would increase As distribution in soluble and readily

exchangeable chemical forms due to the weakened regulatory role of magnetite and its

derivatives.

2.3.3 Arsenic speciation in solution regulated by microbial processes and redox

conditions

Redox conditions (Eh) in soil or tailings can significantly influence As transformation and

speciation in the solution phase (Table 2-4). The spatial heterogeneity of redox conditions in

the rhizosphere can influence the composition and relative proportions of As species in soil

solution. Under semi-arid climatic conditions, oxidising conditions prevail in the rhizosphere

of revegetated plants. Under these kinds of conditions, inorganic As species in oxidised

valency (As5+

) would account for the major proportion of plant available As pool in the

tailings under revegetation. Under reducing conditions (e.g. -200 mV), the predominant

species in the soil solution is arsenite, which may be further converted into organic As forms

through microbial methylation, but under oxidising conditions, arsenate becomes the major

19

arsenic species (Ascar et al. 2008b). However, the redox factor plays a minor role in the

transformation between arsenate and arsenite under Fe-rich conditions (Al-Abed et al. 2007).

This may further highlight the critical influence of magnetite contents on As distribution and

speciation in the Cu-tailings subject to the altered mineral processing to recover magnetite for

economic purposes at the EHM. The distribution of As in the solution phase and inorganic

forms is closely related to As uptake by plants used to phytostabilise the tailings.

Table 2-4: Effect of Eh-pH conditions on As speciation in the aqueous environment (Schnoor

1996)

As speciation H3AsO4 HAsS2 As2S3 H2AsO4- H2AsO3

2- HAsO3

2-

Eh (mv) 700 0 -100 100 -300 -500

pH 1 2 5 7 10 13

In the solution phase, microorganisms including gram-positive bacteria and archaea play

important roles in As speciation by converting inorganic forms into the organic, including

monomethylarsonic acid MMA dimethylarsinic acid DMA, trimethylarsine oxide TAMO and

phenyl arsenic (Cullen and Reimer 1989; Zhang and Selim 2008). The demethylation process

to convert organic As into inorganic forms can be mediated by microbes stimulated by

cellulose, which can degrade arsenobetaine (AsB) to other organic forms of As such as

TMAO, and further DMA and MMA, and finally the inorganic As species (Gao and Burau

1997; Huang et al. 2007). The contribution of organic As species to the total As in soils and

plants is relatively small, compared to inorganic As (Huang and Matzner 2006; Takamatsu et

al. 1982). The organic As (pentavalent) forms are generally less toxic to humans and animals

than the inorganic forms including arsenite and arsenate (Vaughan, 2006). Nevertheless, the

toxicity of trivalent organic As (e.g. MMA (III) (CH3AsH2)) can be much greater than

inorganic As (Zhang and Selim 2008). It is unclear what effects organic matter amendment

will have on the structure and functions of oxidising and acidifying bacteria in the tailings. It

is plausible to hypothesise that the removal of magnetite would increase speciation of the As

into organic forms in the tailings amended with organic matter, because of the significantly

reduced Fe-minerals and associated As-sorbing capacity in the solid phase, compared to the

tailings of high magnetite content.

20

2.4 Arsenic uptake, transport and distribution in plants

2.4.1 Uptake mechanisms in roots influencing factors

Plant uptake of As is closely related to As speciation in the rhizophere and inorganic As

species such as arsenate and arsenite in solution phase are predominantly taken up by plant

roots. However, plant roots may also absorb organic As species (Meharg and Hartley-

Whitaker 2002; Zhao et al. 2009). In natural soils, less than one tenth of As is taken up by

plants which may contain approximately < 1 mg As kg-1

dry weight in the tissue, while

tolerant plants species in As-contaminated soils or tailings soils take up and accumulate

greater amounts of As in the shoot and root than the background (Bañuelos and Ajwa 1999;

Cullen and Reimer 1989). For example, the hyper accumulator P.vittata has been found to

contain more than 1000 mg kg-1

dry weight (Ma et al. 2001). Arsenic uptake and

accumulation in plants vary among plant species, exhibiting responses of As exclusion to

hyper accumulation in the plants (Caille et al. 2004; Zhao et al. 2010).

The uptake of arsenate in root cells shares the same uptake pathway with phosphate (Zhao et

al. 2009) due to their similar chemical properties in bulk soil solution (Asher and Reay 1979).

Arsenate may be associated with 2 protons in the form of H2AsO4-, which is absorbed

through phosphate (Pi) transporters located at the epidermis in the roots (Smith et al. 2010).

In the cytoplasm of root cells, the majority of arsenate, once absorbed, can be rapidly reduced

to arsenite by arsenate reductases before being transported into the xylem by Pi-transporters.

Some of the arsenite may be either converted to organic As species, while only a small

amount of arsenate can also be transported to xylem by Pi transporters (Smith et al. 2010).

The competition between phosphate and arsenate in the root uptake process can be variable,

depending on the growth conditions imposed (Table 2-5). Although arsenate shares the same

uptake pathway as phosphate, the high Pi affinity uptake system preferentially absorbs

phosphate, with increasing phosphate availability, leading to the antagonistic interactions

against arsenate uptake (Ullricheberius et al. 1989). On the other hand, arsenate can be

absorbed through the low affinity system in As tolerant plants such as HolcuslanatusL.,

which has no high affinity Pi uptake pathway without being affected by increasing phosphate

supply (Meharg and Macnair 1990).

Arsenite may become predominant under persistent reducing conditions, which is mainly

present in the form of undissociated arsenous acid (pKas is 9.2, 12.1 and 13.4) at normal pH

(<8.0) such as in paddy soils (Marin et al. 1993; Takahashi et al. 2004; Xu et al. 2008).

21

Arsenite concentration in paddy soils ranges from 0.01 to 3 µM, which is considerably higher

than arsenate in natural soils. The reducing environment can increase arsenite in soil pore

water and its bioavailability for rice (Xu et al. 2008). In contrast to arsenate, plant roots

absorb arsenite in the form of undissociated arsenous acid (H3AsO3), which is affected by

glycerol and antimonite rather than phosphate (Zhao et al. 2009). A glycerol-transporting

channel in S. cerevisiae (baker’s yeast) was found to transport As (III) (pKa=9.2) into the cell

(Quaghebeur and Rengel 2004). In plant roots, some aquaporins in the cytoplasm of the

epidermis and cortex cells can transport the neutral molecules such as As (III), glycerol and

Si from root cells to the xylem (Verbruggen et al. 2009).

Table 2-5: Influences of phosphate on arsenate uptake in different species. The experimental

information has been extracted from the literature

Plant species Experimental

conditions

Impact on As (V)

uptake

References

Barley hydroponic decreased (Asher and Reay

1979)

Holcus lanatus L. hydroponic decreased (Meharg and Macnair

1990)

Silene vulgaris hydroponic decreased (Sneller et al. 1999)

Ryegrass pot increased (Jiang and Singh

1994)

Holcus lanatus L. pot increased (Quaghebeur and

Rengel 2004)

Corn pot increased (Jacobs and Keeney

1970)

Tobacco field increased (Small and McCants

1962)

Plants roots absorb organic As species less effectively than the inorganic As species

including arsenate and arsenite (Burlo et al. 1999; Carbonell-Barrachina et al. 1998;

22

Carbonell-Barrachina et al. 1999; Marin et al. 1992). However, some plant species may

concentrate a larger proportion of organic As than others (Table 2-6). A small amount of

organic As chemical compounds such as MMA, DMA, TMAO also exist in plants. Through

examining the As species in 46 different species of plants, Raab (2007) observed that the

plants could accumulate approximately half the MMA and less than one third of the DMA

when compared with As (V) concentration. MMA is the primary product in the methylation

process, which can be further converted into DMA (Zhao et al. 2010). Bentley (2002) found

that Challenger pathway processing in fungi and bacteria controlled arsenic methylation in

plants. The mechanism of concentrating MMA and DMA by plants remains unclear.

Table 2-6: Organic arsenic accumulated in plant tissues.

Species Concentrations of organic As (%) References

Lycopersicon

esculentum

Less than 1% (Xu et al. 2007)

Hordeum

jubatum

Approximately 3.6% (Koch et al. 1999)

Trifolium

pratense

MMA 35% DMA 24% (Geiszinger et al. 2002)

It is generally known that the rhizosphere microenvironment is different from bulk soil due to

the pH conditions and microbial activity at root surface. The pH in the rhizosphere is results

from rhizosphere Pi supply and root Pi uptake, coupled exudation of organic acids, and CO2

release by root respiration (Meharg 2012). Some arsenite can be converted to arsenate in the

rhizosphere, which results from oxygen release from roots such as rice roots. Arsenate is

easily adsorbed by Fe hydroxide/oxyhydroxide (Fendorf et al. 2007), which present in Fe

plaque at the root surfaces resulting from the oxidation of ferrous iron into ferrihydride (Blute

et al. 2004; Hansel et al. 2002). This portion of As adsorbed by the Fe-plaque at root surface

is the reason behind the overestimation of total As in the root. In addition, arsenite may be

exuded from roots, which can be oxidized into arsenate in the rhizosphere (Logoteta et al.

2009; Zhao et al. 2010).

23

2.4.2 Arsenic transport and distribution in plants

Following absorption by roots, the majority of the arsenate is reduced into arsenite in root

cells by arsenate reductase; arsenite is thus the predominant form of As transported in the

xylem (Pickering et al. 2000; Raab et al. 2005; Xu et al. 2007; Zhao et al. 2010). The group

of thiol-rich peptides--glutathione (GSH) and phytochelatins (PCS) are associated with

trivalent As species in the xylem, grain (e.g. rice), roots and shoots (Oscarson et al. 1981; Xu

et al. 2007; Zhao et al. 2010).

In leaves, arsenite is transformed to monomethylarsonate (MMA) and dimethylarsinate (Wu

et al. 2002). MMA is the primary organic product in methylation processing, while DMA and

TMAO appears in leaves after the procedure due to the reduction step. The final product is

TMA (trimethylarine), which will transpire out of leaves. Organic methylated arsenic

speciation is predominant in wheat leaves (Zhu et al. 2006).

The mobility of As species in plants is much lower than that of phosphorus during the

transport process from root to shoot (Zhao et al. 2003). This is because phosphorus is an

essential element in plants. Plants need to accumulate sufficient phosphorus for basic

physiological activities such as photosynthesis, respiration, cell division and energy

metabolism in every growth stage and in different organisms. In contrast, arsenic is

nonessential in plant growth and development and indicates a negative function for plants.

Therefore, arsenic is more likely to remain in roots based on the tolerance strategy; different

plant species have different capacity to accumulate As in roots and shoots (Table 2-7).

24

Table 2-7: Arsenic distribution in plant tissues of different plant species

Plant species Experiment

conditions

As

concentration in

roots (mg kg-1

)

As aboveground

concentration

(mg kg-1

)

References

Maize Sand (As 240

mg kg-1

)

227 leaves 11 (Gulz et al.

2005) stem 9

grain 0.1

Tomato Soil and solution

(As 2 mg L-1

)

125.8 leaves 2.15 (Barrachina et

al. 1995) stem 2.29

fruit 0.138

Ryegrass Sand (As 240

mg kg-1

)

255 leaves 11 (Gulz et al.

2005) stem 16

grain ND

Rice Solution (As 1.5

mg L-1

)

248 leaves 14 (Smith et al.

2008) stem 17

grain 0.87

Spanish Flora

(Native species)

Soil (As 632 mg

kg-1

)

N.D. 0.65-2.52 (Moreno-

Jiménez et al.

2010)

2.4.3 Species diversity in As uptake and accumulation

Different plant species may exhibit a range of As accumulation behaviours in both natural

and cultivated growth conditions, which can be categorised into excluder, accumulator and

hyper accumulator species, depending on the uptake, transport and accumulation behaviour in

the plants (Baker 1981; Prasad 2008). The Bioaccumulation factors (the ratio of As

concentration in plant tissue to As in soil or water) are commonly calculated to separate

excluders, accumulators and hyper accumulators, which have BF values of <1, 1-10, and > 10,

respectively (Baker 1981; Ma et al. 2001) (Table 2-8). In nature, some As hyper accumulator

plant species have been found and well-studied (Kachenko et al. 2007; Shoji et al. 2008;

25

Srivastava et al. 2005; Zheng et al. 2008). These species indicate a strong affinity for arsenic

uptake and translocation, such as the fern species of the Pteris family (Ma et al. 2001). In

hyper-accumulators, arsenite is accumulated in leaves more than roots (Pickering et al. 2006).

Table 2-8: Comparison of arsenic accumulation among different plant species grown in As-

contaminated soils

Plant group Plant species As in

substrate

As in plants Bioaccumul

ation factor

References

Excluder Bidens pilosa

125 mg kg-1

13.5 mg kg-1

0.11 (Sun et al.

2009)

Yellow lupines 10 mg kg-1

1.5 mg kg-1

0.15 (Dary et al.

2010)

Accumulator Ryegrass(Lolium

perenne)

58 mg kg-1

102 mg kg-1

1.76 (Hartley and

Lepp 2008)

A. blitoides S.

Watson

78 mg kg-1

114 mg kg-1

1.46 (Del Rio et

al. 2002)

Hyper

accumulator

Pteris vittata 300 mg kg-1

6158 mg kg-1

20.1 (Lombi et

al. 2002)

In addition, in vascular plants, arsenic transportation from root to shoot is more effective in

dicots than monocots (Bondada 2003; Otte et al. 1993) (Table 2-9).Therefore, in native grass,

As concentration in dicot legume species is more effective than that in monocot poaceae

species.

26

Table 2-9: Examples of arsenic transport from roots to shoots of different plant species

Plant

species

Dicot or

monocot

Experiment

conditions

Arsenic in

root (mg

kg-1

d. wt)

Arsenic in

shoot (mg

kg-1

d. wt)

Ratio of

As[shoot]

to As[root]

References

Winter

wheat

monocot Solution

Arsenic 20

mg L-1

4100 75 0.018 (Liu et al.

2008)

Maize monocot Soil Arsenic

100 mg kg-1

220 3.2 0.014 (Yu et al.

2010)

Rice monocot Solution,

10µM

Na3AsO4

1200 50 0.042 (Geng et al.

2005)

Sunflower dicot Soil Arsenic

620 mg kg-1

347 74.8 0.22 (Ultra et al.

2007)

Radish dicot Solution

Arsenic 2

mg L-1

35.5 9.6 0.27 (Smith et al.

2009)

Mesquite

(Prosopis

juliflora)

dicot Solution

Arsenic 10

mg L-1

980 450 0.46 (Mokgalaka-

Matlala et

al. 2008)

2.5 Impacts of tailings amendment on As availability and plant uptake

2.5.1 Organic amendment

Soil organic matter from plant and animal litters and some root exudates is the source of

nitrogen (N), sulphur (S), phosphorus (P) and other elements (Sharma et al. 2010). Organic

matter improves soil microbial activity and the formation of soil aggregates (Noyd et al. 1996;

27

Wright and Upadhyaya 1998). As a result, the addition of organic matter is a common

practice in tailings revegetation.

In mine tailings with a high total As concentration, the As bio-availability can be increased

by the addition of organic matter (Bauer and Blodau 2006; Meunier et al. 2011). Fulvic acid

and humic acid from dissolved organic matter can form stable complexes on the mineral

surfaces and chelating As species that bond to Fe oxides, alumina, quartz or kaolinite (Grafe

et al. 2001; Grafe et al. 2002; Kaiser et al. 1997; Xu et al. 1991). They can increase As

mobility by displacing the bonding arsenate and arsenite from the Fe oxide surfaces (Redman

et al. 2002). In another experiment, Wang and Mulligan (2009a) found that As mobility could

be enhanced by natural organic matter, forming aqueous complexes.

Organic matter amendment plays an important role in inorganic chemical speciation of As by

converting inorganic forms into organic As species (Sohrin et al. 1997; Turpeinen et al. 1999).

As a result, the pore water toxicity level could be decreased due to the high toxic inorganic

As species being transformed to the less toxic organic As species. Therefore, these

microorganisms play an important role on amending soil and water. For example, Jahan et al.

(2006) used the arsenic-resistance bacteria (Scenedesmus abundans) collected from waste

plants to remove nearly 70% arsenic in groundwater.

In the phytoremediation process, plant roots can accumulate less inorganic As for

transporting into shoots by adding organic matter due to the transformation of inorganic As,

therefore reducing the risk for livestock and human exposure to these plants.

2.5.2 Inorganic amendment impacts

Arsenic chemical properties are similar to phosphorus, especially from the aspects of arsenate

and phosphate. This leads to competition between arsenate and phosphate in sorption

processing (Gunes et al. 2009; Smith et al. 2002), which means additional phosphate inhibits

As sorption in As rich soil. Phosphorous also shows an affinity for Fe oxides and

oxyhydroxides in soil and groundwater (Mucci et al. 2000; Tejedor-Tejedor and Anderson

1990). In iron amended soils without As, phosphorous concentration can be reduced in soil

pore water, which causes nutrient depletion, however, in liquid solution, phosphorus

concentration can be increased around the rhizosphere due to the adsorption of iron cations

present in the rhizosphere (Zhang et al. 1999). Therefore, the competition reduces the

28

sorption of As and P on the iron surface (Gao and Mucci 2001) and increases the release of

bio-accessible arsenate and phosphate from the surface of tailings into pore water.

2.6 Organic amendment and phytostabilization of Cu mine tailings

Phytostabilization with native plant species is a preferred closure option strategy of tailings

facilities. This requires the reconstruction of functional root zones as a pre-requisite, which

can be achieved by adding amending materials such as organic matter and lime and

engineering the profile (Huang et al. 2011; Li and Huang 2014). To identify effective

revegetation options at EHM, a long-term semi-controlled trial has been in place since 2010,

in which the old tailings (with high magnetite content) were amended with dry hay (up to 20%

volume) and sown with native grasses including Red Flinders Grass, Bull Mitchell, Curly

Mitchell and Barley Mitchell (Huang et al. 2011). Initial results have indicated the

establishment of primary root zone functions due to improved physical and hydraulic

properties from the organic matter amendments (Huang et al. 2011). Considering the

availability of organic matter in Queensland, sugarcane residue is abundantly available in

central Queensland and may have similar effects to hay, as both contain high proportions of

labile carbon. In addition, there are tonnes of waste wood (pine pellets) on site which can be

pyrolyzed into biochar and stocked up for tailings amendment at the time of mine closure.

Biomass-based biochar can not only provide physical improvement due to the stable and

porous carbon structure in the biochar, but have chemically functional surfaces for adsorbing

heavy metals (Cu, Zn, Pb, Ni and Fe etc.) and metalloids (As) in the tailings and

contaminated soil (Sneath et al. 2013; Uchimiya et al. 2012). Therefore, biochar is a cost-

effective and eco-friendly option. The direct revegetation trial has been conducted at the

EHM-tailings (high magnetite) storage facility since 2010 (Huang et al. 2011).

In response to the proposed processing technology to recover magnetite from the Cu-tailings

at EHM, questions have been raised whether the reduction of magnetite from the tailings

could change the chemical forms of As favouring exchangeable and soluble forms, thus

increasing As risks in the tailings under revegetation. As a result, the primary aim of the

present study investigated if the tailings of low magnetite content would have different As

distribution among different chemical forms from those in the high magnetite tailings. This

was followed by a detailed examination of As speciation in the pore water of the tailings with

contrasting magnetite contents and organic matter amendments, and As accumulation by the

native grass (Red Flinders grass). The findings contribute to the understanding of As risks in

29

the tailings and seepage after magnetite recovery, which forms the basis for developing

corresponding tailings management options in the closure plan at Ernest Henry Mine.

On the basis of the current understanding of the closely coupled relationship between Fe-

minerals and As adsorption, and OM-regulated microbial activities in As speciation, the

present project has investigated the following specific hypotheses:

1. The presence of high magnetite contents would increase As adsorption due to the

increased adsorption sites from oxidised Fe coating magnetite particle surfaces,

compared to low magnetite tailings. It is therefore expected that magnetite would

regulate the distribution of As into the soluble pool and higher proportions of As

would be distributed in pore water of the low magnetite tailings than the high

magnetite tailings.

2. Due to the adsorption of As by Fe-oxyhydroxides in high magnetite tailings, less

soluble As is available for microbe-mediated transformation from inorganic into

organic species.

3. Based on the As re-adsorption onto the surfaces of magnetite particles, plant available

As in pore water of low magnetite tailings would be greater than the high magnetite

tailings.

The current project consisted of laboratory experiments to analyse different chemical forms

of As in relation to physicochemical properties and glasshouse experiments in which the

tailings of contrasting magnetite contents were amended with organic matter treatments

(sugarcane residue and pine biochar) and planted with native grass species.

30

Chapter 3 Altered arsenic distribution in copper tailings of contrasting magnetite

content and under organic matter amendment

3.0 Introduction

Arsenic is an abundant trace element in Cu-tailings originating from As-bearing sulphides

and sulphosalts, including arsenopyrite (FeAsS), cobaltite (CoAsS), enargite (Cu3AsS4) and

tennantite (Cu12As4S13) (Dold and Fontboté 2001; O'Day 2006). The As contained in the

minerals can be released into the pore water during the process of oxidation and dissolution.

The As can then in turn be strongly adsorbed by secondary minerals, e.g., Al/Fe/Mn oxides

and oxyhydroxides (particularly in amorphous phase), under suitable biogeochemical

conditions (Drahota and Filippi 2009; Kumpiene et al. 2012; Wang and Mulligan 2008). As a

result, the As distribution in different forms in the tailings may be altered by changes in ore

processing technology that removes As-adsorbing minerals, for example, reprocessing to

extract magnetite in the Ernest Henry Mine (EHM) tailings. The solubility and phyto-

availability of As in mine tailings are closely related to As chemical forms in the solid phase

tailings. These induced changes may increase potential environmental risks in both As

mobility in the seepage water and plant As uptake in revegetated plant ecosystems on the

tailings impoundment.

Characterisation of As distribution in the solid phase has been commonly used to assess the

solubility and phyto-availability of As in contaminated soil and mine tailings (Kumpiene et al.

2012; Niazi et al. 2011). On the basis of chemical extractability, As forms in the solid phase

may be partitioned into water soluble, phosphate exchangeable, organic matter complexes,

and Fe/Al/Mn/Ca bound minerals (Kim et al. 2003; Quazi et al. 2010; Wenzel et al. 2001).

The As forms associated with secondary Fe/Al/Mn minerals are more stable than the soluble-

As and exchangeable-As, but still they can be dissolved into pore water under persistent

reducing conditions and with the aid of Fe-reducing bacteria (Bauer and Blodau 2006; Sarkar

and Datta 2004a; Sarkar and Datta 2004b). In contrast, the phosphate exchangeable-As can

be released into the pore water when exposed to elevated levels of phosphate in the solution

as phosphate can compete with arsenate for sorption sites of the Fe oxides and Fe-

oxyhydroxides (Cao et al. 2003; Lenoble et al. 2005; Manning and Goldberg 1996).

Therefore, sequential As fractionation of tailings and contaminated soil/sediments has been

commonly applied to understand the potential mobility of As in soils and sediments, in

31

relation to the mineralogy and biogeochemical conditions concerned (Palumbo-Roe et al.

2007).

Iron minerals play a significant role in As immobilisation in tailings through Fe-dissolution

and precipitation of Fe-oxyhydroxides at surfaces of crystalline Fe-minerals and other

minerals under suitable redox conditions. Ferrihydrite or hydrous ferric oxides have a higher

surface area and affinity for phosphate than magnetite or lepidocrocite, adsorbing 10 times

more Pi than the latter (Barber 2002). Previous studies indicated that magnetite particles

could adsorb As on their surfaces (Giménez et al. 2007; Yean et al. 2005). Furthermore,

bioreduction or biomineralisation of Fe-minerals (e.g. magnetite, hematite) can be facilitated

by dissimilatory metal reducing bacteria which can transform Fe (III) into Fe (II), leading to

the formation of many secondary Fe oxides and oxyhydroxides with a high affinity for As (V)

and Pi sorption (O’Loughlin et al. 2013; Zachara et al. 2002). In the EHM Cu tailings, the

economically-driven recovery of magnetite decreased magnetite content from as high as 29%

to as low as 3% in the Cu tailings; the magnetite removal is the physical process using

grinding classification and magnetite separation (Davey KJ 2008) (Figure 3-1). Under

revegetation, Fe-oxyhydroxides may be formed on the surfaces of magnetite particles and

adsorb soluble As from the pore water in the Cu tailings. As a result, we hypothesised that the

low magnetite tailings content (LM) would change As chemical forms in the solid phase and

favour its distribution in readily exchangeable forms (such as water-soluble, non-specifically

sorbed and specifically sorbed fractions) in comparison to the high magnetite tailings

contents (HM).

32

Figure 3-1: The process of magnetite removal in EHM tailings and flotation test

(Davey KJ 2008)

Non-Magnetite Fraction

p80 = 100 µm

Ball mill

Magnetite

Separator

Magnetite

Hydrocyclone

Plant tailings

p80 = 200 µm

Hydrocyclone Underflow

Ball mill

Hydrocyclone

p80 = 100 µm

Hydrocyclone

Tower Mill

p80 = 38 µm

p80 = 38 µm

Magnetite

Separator

Magnetic Conc Tailing

Hydrocyclone

Overflow

d50 = 25 µm

33

Organic matter amendment is often required to improve physical and biogeochemical

conditions for improving plant growth for phytostabilization purposes (Huang et al. 2012).

Organic matter and its derivatives from its decomposition can not only directly influence As

chemical forms through competing functional groups (such as phenolic, carboxyl, hydroxyls,

etc.), but also enhance microbes-mediated processes to transform Fe-primary minerals into

secondary minerals such as Fe-oxyhydroxides (Behrends and Van Cappellen 2007; Redman

et al. 2002). Iron reducing bacteria play a significant role in the transformation of both

amorphous and crystalline iron minerals, which may affect the As distribution on the Fe

phase (Cummings et al. 1999). For example, Coker (2006) found that ferrihydrite could be

reduced by the reducing microbe Geobacter sulfurreducens and form magnetite at its surface.

Dong (2000) observed that magnetite could be converted to vivianite or siderite by the

reducing bacteria Shewanella putrefaciens. Therefore, organic matter amendment may

increase the formation of the As-adsorbing minerals Fe-oxyhydroxides, which coat the

surfaces of magnetite particles and thus favour the As forms associated with secondary Fe

minerals (e.g. amorphous Fe oxides/oxyhydroxides).

As a result, the primary objectives of the present chapter are to investigate (1) whether the

high magnetite content provides high As adsorption capacity in the tailings using batch tests;

(2) whether the reduction of the magnetite in the tailings favours As distribution in non-

specifically adsorbed forms; and (3) whether the addition of organic matter increases As

distribution in specifically adsorbed forms in the EHM Cu-tailings of contrasting magnetite

content. Two types of organic matter were used to improve the tailings physical and chemical

properties for plant establishment: sugar cane residues (easily decomposable) and pine

biochar (stable C from high temperature (600-650 °C) pyrolysis). The amended tailings were

incubated in well-watered pots under glasshouse conditions. The tailings (including both high

and low magnetite tailings) were collected from the tailings impoundment area of EHM,

which has ores of copper (chalcopyrite)-gold mineralisation mainly within the magnetite-

biotite-calcite ± pyrite matrix of a pipe-like breccia body (Ryan 1998). The mine is located in

Cloncurry, Northwest Queensland. Arsenic is present in these ores as a waste product in the

copper extraction process, which contains arsenic minerals (mainly cobaltite (CoAsS) and

arsenopyrite) and Fe-oxides (mainly hematite and magnetite) in the tailings (Forsyth 2010).

34

3.1 Materials and Methods

Bulk tailing samples of the high and low magnetite (referred to hereon as HM and LM,

respectively) contents were collected from EHM tailings impoundment area in August 2012

and transported to the laboratory. The tailings samples were air-dried and sieved through a 2

mm sieve prior to physicochemical analysis. Two types of organic matters were used in the

present experiment, including sugarcane mulch residue (SR) (Earth Wise Company, Qld) and

pine biochar (BC) (ANZAC Pty Ltd, pyrolysed at 700 °C). The organic matter samples were

dried at 65 °C, ground and passed through a 2 mm sieve.

The glasshouse incubation experiment aimed to investigate the effects of direct revegetation

treatments on the distribution of As in various chemical forms and alterations in chemical

properties. The detailed experimental design can be found in Chapter 4. Briefly, the

treatments in the HM and LM tailings include: control, SR (5% w/w), SR (5% w/w) + BC

(1%), and SR (5%) + BC (5%). In the amended treatments, Red Flinders grasses (Iseclema

Vagin florum) were planted; and the plants did not survive in the unamended tailings

(control). The addition of SR (5%) as a base treatment in the amended treatments improved

the survival of the grass.

3.1.1 Physicochemical analysis: pH, EC, particle size and total element concentrations

The measurement of pH and electrical conductivity (EC) measurements was made by means

of a bench-top TPS 901-CP meter. The EC and pH were measured in 1:5 soil (or solid)/water

suspensions. Particle size in the tailings samples was measured by the Mastersizer 3000 laser

diffraction particle size analyser (Malvern instruments Ltd) (JKMRC, University of

Queensland). Total concentrations of metal and metalloids and other relevant elements in the

LM and HM tailings were determined by using Inductively Coupled Plasma Optical Emission

Spectroscopy (ICP-OES) (Varian Vista Liberty, Australia) following aqua-regia digestion

using an AIM600 Block Digestion System (Aim Lab Automation Technologies Pty Ltd). The

digestion method has been described by Wheal (2011). Briefly, aliquots of 0.4 g air-dried

tailing samples were transferred into 100 ml pyre glass digestion tubes, to which

approximately 10 ml aqua regia (HNO3: HCl 1:3, made fresh) were added in each tube. The

sample-acid mixtures were digested at 140 °C for a minimum of 3 hours based on

preliminary investigation. Quantitative XRD analysis for LM and HM tailings was conducted

at United Minerals Services (NSW Australia) using the method described by Ortiz (2009).

35

3.1.2 Fe Mn Al extraction

The air-dried tailings samples (including those amended with organic matter for 4 weeks and

the unamended control) were extracted for acid-oxalate (Ox) and dithionite-citrate-

bicarbonate (DCB) extractable Fe/Al/Mn oxyhydroxide (Rayment and Lyons 2010). The

acid-oxalate-extractable Fe, Al, and Mn originate from amorphous and poorly crystalline

minerals including ferrihydrite, allophone, imogolite and minerals containing Fe2+

such as

magnetite. Dithionite-citrate-bicarbonate (DCB) extractable Al and Fe originate from both

crystalline and non-crystalline Fe oxides and Al substituted in crystalline Fe oxides in soils,

including those from hematite and goethite and amorphous Fe and Al. Similar extraction

protocols were used for Mn (Mahaney et al. 1994).

Oxalate extracting reagents were prepared using ammonium oxalate [(COONH4)2.H2O] and

oxalic acid [(COOH)2.2 H2O] with pH buffered at 3.0 ± 0.05. Samples were LM and HM

extracted in a mixture of tailings-solution at 1:100 (wt:v) for 4 h at 25 °C in the dark (i.e.

bottles were covered with aluminum foil to prevent exposure to the light), on an end-over-end

shaker. After mixing, 50 ml supernatant was collected from each bottle and then centrifuged

at 3600 g for 10 minutes, and further filtered through 0.45 µm filter paper. The filtered

solution was acidified with 100 µl 70% HNO3 for each 10 ml volume prior to ICP-OES

analysis.

The citrate-dithionite solution consisted of sodium citrate (Na3C4H5O7.2H2O) and sodium

dithionite (Na2S2O4) (Sigma-Aldrich chemical, AR grade). Aliquots of 1g EHM LM and HM

air-dried tailing samples were weighed into 200 ml plastic bottles, to which 1g sodium

dithionite was added per bottle, together with 50 ml extracting solution. The sample bottles

were shaken in an end-over-end shaker for 16 h at 25°C. After shaking, distilled water (50

ml) was added to each bottle for initial dilution and 5-6 drops superfloc solution were added.

Samples were then shaken vigorously for another 10 seconds. Aliquots of 50 ml extracts were

transferred into 50 ml falcon tubes, which were centrifuged at 3600 g for 10 minutes. The

supernatant solution was filtered through 0.45 µm filter paper and acidified with 100 µl 70%

HNO3 prior to ICP-OES analysis.

3.1.3 Arsenate adsorption

Arsenate adsorption isotherms were established to compare the relative As adsorption

capacity in the tailings between high and low magnetite contents without organic matter

amendment, in 0.01 M NaCl solution (as the background electrolyte) at pH 7.0, which

36

contained a range of As concentrations. The arsenate solutions were prepared by dissolving

sodium arsenate dibasic heptahydrate (Na2HAsO4. 7H2O Sigma-Aldrich Chemical, AR grade)

in 0.01 M NaCl. The nominal As (V) concentration (mg L-1

) were 0 (control), 1, 5, 10, 20,

and 50. Aliquots of 1 g air-dried tailings (< 2mm) were weighed in 50 ml Falcon tubes, to

which 20 ml arsenate solution was added. Each As (V) concentration was replicated three

times. The tailing-As solution mixtures in the tubes were shaken for 24 h at room temperature

(about 25 °C) on an end-over-end shaker. At the end of the shaking process, the mixtures

were centrifuged at 3600 g for 10 minutes and the supernatants were further filtered through a

0.45 µm filter, prior to ICP-OES analysis. Total As concentrations in the treatment solutions

were quantified before and after the adsorption process by means of ICP-OES.

3.1.4 Arsenic fractionation

The HM and LM tailings (including the amended and unamended) were fractionated for As

chemical forms according to Wenzel (2001) (Table 3-1). Extracting reagents were made up

in Millipore water, by dissolving (NH4)2SO4, NH4H2PO4, NH4-oxalate, NH4-oxalate and

oxalic acid (Sigma-Aldrich Chemical, AR grade). The amended tailings with organic matter

were collected from a glasshouse experiment (4 weeks) in which native grasses were planted.

Each sample had four replicates. After the sequential extraction steps, the residue tailings

were air-died in a 65 °C oven for 2 days and digested in concentrated nitric acid at 120 °C

with a heating block (AIM600) for total As concentration. The detailed setup and sampling

procedures of the glasshouse experiment with the HM and LM tailings treated with organic

matters can be found in Chapter 4.

37

Table 3-1: Arsenic fractionation method by Wenzel

Fraction step Extracting

chemicals

Conditions Solution

volume

Wash step

Non-specifically

sorbed

(NH4)2SO4

0.05M

25 °C, 4h

shaking

25 ml 10 ml DI water

Specifically-sorbed NH4 H2PO4

0.05M

25 °C, 15-16 h

shaking

25 ml 10 ml DI water

Amorphous Fe(III)-

oxyhydroxide

NH4-oxalate

buffer 0.2 M pH

3.25

25 °C, 4h

shaking in dark

25 ml NH4-oxalate 0.2

M 12.5 ml (pH

3.25) shake in

darkness

Crystalline Fe(III)

oxyhydroxide

NH4-oxalate

buffer 0.2 M and

ascorbic acid 0.1

M pH 3.25

96±3 °C in

water basin, 30

min in bright

25 ml NH4-oxalate 0.2

M 12.5 ml (pH

3.25) shake in

darkness

Residual 70% HNO3 Heating block 10 ml None

3.1.5 Data analysis

The Langmuir model (qe=QmaxbCe/1+bCe) was used to fit the relationship between the

amount of As adsorbed by the tailings and the As concentrations at the end of adsorption test,

from which the adsorption capacity for arsenate was estimated for the two tailings. A two-

way analysis of variance was carried out to evaluate treatment effects and their interactions

(SPSS 20.0, IBM, USA). The differences among the means were compared by using LSD-

0.05. Correlation analysis was performed to evaluate the relationships among various sets of

parameters as indicated in each table or figure.

3.2 Results

3.2.1 Physicochemical properties

The two tailings had neutral pH conditions, while the BC appeared slightly alkaline (Table 3-

2). The HM tailings (1.98 ms cm-1

) were slightly more saline than the LM (1.69 ms cm-1

).

38

The particle size distribution of the two tailings was similar (Figure 3-1). The fraction of

coarse silt accounted for the largest proportion among the size groups. The HM tailings

contained a slightly higher proportion of coarse silt than the LM. This may have resulted

from the further grinding of the tailings for magnetite recovery and the removal of the size

fraction of magnetite particles. The total concentrations of Fe, As, Al and Mn in the LM

tailings were approximately 31%, 37%, 5.2% and 35% of those in the HM, respectively. In

terms of the Quantitative XRD analysis, the crystalline Fe minerals composition (magnetite,

hematite and pyrite) in HM tailings was higher than that of LM tailings, with approximately

12% magnetite, 2.5 % hematite and 5.6% pyrite in HM tailings compared to that of 4.6%, 1.3%

and 4.3% in LM tailings. The total Fe, Al and Mn concentrations in the organic materials

were significantly lower than those in the tailing materials. The levels of total As in SR and

BC samples were not detected.

Figure 3-2: Particle size distribution in EHM tailings (clay: 0-2 µm; fine silt: 2-6.3 µm;

medium silt: 6.3-20 µm; coarse silt: 20-63 µm; sand: >63 µm)

0%

10%

20%

30%

40%

50%

Clay Fine silt Medium silt Coarse silt Sand

LM tailings

HM tailings

39

Table 3-2: Background physicochemical properties of EHM tailings SR and BC (air-dry

weight) used in the experiment, including pH water, EC, total element concentrations and

crystalline Fe minerals composition (Quantitative XRD)

Measurements LM tailings HM tailings SR BC

pH 7.1±0.1 7.1±0.2 6.5±0.1 7.8±0.2

EC (ms cm-1

) 1.7±0.1 1.98±0.02 2.4±0.1 2.2±0.1

Total As (mg kg-1

) 144.3±4.6 387.4±13.1 0 0

Total Fe (g kg-1

) 83.2±1.5 109.4±1.4 2.4±0.2 1.3±0.2

Magnetite (%) 4.6 12 0 0

Hematite (%) 1.3 2.5 0 0

Pyrite (%) 4.3 5.6 0 0

Total Al (g kg-1

) 9.6±0.2 10.1±0.6 2.4±0.1 1.1±0.1

Total Mn (g kg-1

) 2.3±0.1 3.1±0.2 0.3±0.01 0.06±0.01

The values are the means of three replicate analyses with the “±” indicating standard

deviation.

3.2.2 Distribution of Fe/Al/Mn oxyhydroxides

The citrate-dithionite extractable Fe (crystalline) in the LM tailings was only 43% of the HM,

but there was no significant difference in ammonium-oxalate extractable Fe (Amorphous)

between LM tailings and HM tailings (Table 3-3). Both amorphous and crystalline Mn

concentrations were lower in the LM tailings than those in HM, which only accounted for 16%

and 46% of the HM tailings. However, both amorphous and crystalline Al in LM tailings

were 82% and 56% higher than those in the HM tailings.

After the 4 weeks of incubation with the treatment of SR and BC and plant growth under

glasshouse conditions, the amorphous and crystalline Fe concentrations in LM tailings were

lower than that in the HM tailings across the amendment. The crystalline Fe concentrations in

the LM and HM tailings were slightly higher than the unamended tailings, with about 8 g kg-1

40

in LM tailings and 12 g kg-1

in HM tailings. In contrast, the amorphous Fe concentrations

increased dramatically in organic matter amended LM and HM tailings, with about 20 g kg-1

in LM tailings and 30 g kg-1

in HM tailings. Across the treatments, HM tailings with 5% BC

indicated the highest amorphous Fe concentration (30 g kg-1

) while HM tailings without BC

indicated the highest crystalline Fe concentration (13 g kg-1

).

Although the organic matter contained Mn (358 mg kg-1

in SR and 64 mg kg-1

in BC), there

was no significant difference in amorphous Mn between amended and unamended treatments

in both LM and HM tailings. In addition, the amorphous Mn distribution in LM and HM

tailings was also similar. The crystalline Mn distributed in the HM tailings was significantly

higher than that in the LM amended with SR and 1% BC. In the LM tailings amended with a

base rate of SR (5%), the crystalline Mn concentration (mg kg-1

air dwt) decreased with

increasing BC rates, which were 1413 in 1% BC and 1326 in 5% BC, respectively, compared

to 1557 in the SR only treatment.

In the LM tailings, the amorphous Al concentrations in the unamended were higher than

those of the amended, which were generally higher than those in the HM tailings (p<0.05).

Among the different BC treatments, there was no significant difference in amorphous Al

concentrations between the LM and HM tailings. In contrast, the crystalline Al concentrations

(mg kg-1

air dwt) in the LM tailings were higher than those in the HM, which were 600-650

and 500-600, respectively.

41

Table 3-3: The distribution of extractable Fe/Mn/Al oxyhydroxides in the LM and HM

tailings, in response to organic matter treatments under direct revegetation with Red Flinders

grass

Tailing samples Fe-d (g

kg-1

)

Fe-ox (g

kg-1

)

Mn-d (mg

kg-1

)

Mn-ox

(mg kg-1

)

Al-d (mg

kg-1

)

Al-ox

(mg kg-1

)

LM only 5±1e 13±1 c 1148±14c 186±2c 307±21f 565±5b

LM+SR 7±1 d 16±2 bc 1073±81c 224±23b 602±24c 653±74ab

LM+SR+1%BC 8±1 cd 23±6 b 1161±169c 304±49b 642±39b 724±158a

LM+SR+5%BC 8±1 c 17±2 bc 1263±116bc 226±16bc 731±22a 605±33ab

HM only 11±1 b 14±2c 1181±32c 516±29a 423±9e 661±19ab

HM+SR 13±1 a 25±1ab 1577±139a 253±4bc 593±48c 489±29bc

HM+SR+1%BC 11±1b 28±2 a 1413±76b 292±20b 503±6d 495±33bc

HM+SR+5%BC 11±1 b 30±2 a 1362±72b 350±37b 624±15bc 482±21bc

The values are the means of three replicate analyses with “±” indicating standard deviation.

The LSD tests to show the differences between treatments for the same parameter and

different letters indicate their significant differences at P<0.05.

Fe-d; Mn-d; Al-d: Dithionite extraction-crystalline Fe Al Mn

Fe-ox; Mn-ox; Al-ox: Oxalate extraction-amorphous Fe Al Mn

Note: Oxalate (pH=3.0) extractable Fe and Al include amorphous Fe and Al and organic-

complex Fe and Al.

3.2.3 Arsenate adsorption by EHM tailings

Both the LM and HM tailings adsorbed arsenate, but the adsorption capacity of the former

was lower than the latter (Figure 3-2). Based on the Langmuir model (LM tailings:

Qmax=153.5 b=0.87 R2=0.9259; HM tailings: Qmax=241.2 b=1.41 R

2=0.9672), the maximum

As (V) adsorption in LM tailings was 153.5 mg kg-1

air dwt compared to 241.2 mg kg-1

in the

42

HM. In addition, Fe concentrations dissolved out of the sample during the adsorption tests

were measured, which were <20 µg L-1

in each samples (40 ml solution).

Figure 3-3: Arsenate adsorption isotherms in EHM tailings which were fitted with the

Langmuir model

3.2.4 Arsenic fractionation

By comparing the main effects of treatment factors and their interactions, non-specifically

sorbed As was significantly elevated by the addition of SR, rather than the tailings type.

There were no interactions between the two main factors (magnetite content and SR residue)

(Table 3-4). However, the other fractions including specifically sorbed As, arsenic sorbed,

amorphous Fe and Al oxides and residual As were significantly altered by both the tailings

type (magnetite) and SR treatments; the effects of the SR treatment appeared to be dependent

on the tailings type. In contrast, in the basal SR amendment, the addition of BC did not

significantly alter the As distribution among the fractions and the main effects on the As

chemical form distribution were caused by the magnetite contents in the LM and HM tailings

(Table 3-6).

The distribution of As forms in the solid phase differed significantly between the LM and

HM tailings without any amendments and revegetation treatments (Table 3-4). Non-

specifically sorbed As was below the detection limit in both LM and HM tailings. The main

43

difference was associated with the specifically sorbed As (exchangeable) fraction, in which

As concentration in the HM tailings was 20% higher than those in the LM tailings. The

concentration of As associated with amorphous Fe/Al oxides in the LM tailings was 11 mg

kg-1

air dwt, but was 113 mg kg-1

air dwt in the HM. The As concentrations in the residue

fraction of the LM and HM tailings was 117 and 198 mg kg-1

air dwt, respectively, which

accounted for the majority of the As in the tailings.

The addition of organic matter in the tailings altered the distribution of As among the

fractions or chemical forms (Table 3-4 and 3-5). The immediately and potentially available

fractions of As were significantly elevated in the LM tailings amended with SR and BC under

direct revegetation, despite the significant reduction of total As in the LM. In the LM tailings,

the distribution of As in the specifically sorbed fraction and the fraction sorbed by amorphous

and poorly crystalline hydrous oxides of Fe and Al were consistently increased by the

amendment of SR, resulting in reduced distribution in the residue fraction. In contrast,

considerably less As was distributed in the specifically sorbed fraction and As sorbed by

amorphous and poorly crystalline hydrous oxides of Fe and Al in the HM tailings subject to

the direct revegetation treatments led to increasing As distribution (by about 25%) in the

residue fraction, compared to the HM tailings without any treatments.

From the correlation analysis, we found that the non-specifically sorbed As was significantly

and positively related to Fe-ox, Al-ox and Mn-d concentrations in the tailings including

amended and unamended LM and HM tailings (Table 3-7). The specifically sorbed As was

only positively related to Mn-ox concentrations. The As sorbed by amorphous Fe and Al

oxides was positively related to tailings pH and concentrations of Fe-d and Mn-ox. The

residual As concentrations were positively related to pH, Fe-d and Fe-ox and Mn-d, but

negatively to Al-ox (Table 3-7). The non-specifically sorbed As was negatively and

significantly (P < 0.05) correlated with the specifically sorbed As fraction (Table 3-8), while

specifically sorbed As was positively and significantly (P<0.01) correlated with the As

sorbed by amorphous Fe/Al oxides, which was also positively correlated with the residual As

(Table 3-7).

44

Table 3-4: Arsenic distribution among the chemical forms in LM and HM tailings (organic

matter amended and unamended tailings). The amended tailings were incubated in a well-

watered status under glasshouse conditions for four weeks, in which Red Flinders grass was

grown

As fractionation Non-

specifically

sorbed

Specifically

sorbed

Sorbed by

amorphous

Fe & Al

oxides

Sorbed by

crystalline

Fe & Al

oxides

Residual Total As

Treatment mg As kg-1

LM only N.D 1±0.2 d 11±0.4 d N.D 117±5 d 144±4 d

LM+SR 1±0.7 a 4±0.5 c 16±1 cd N.D 92±8 d 122±3 e

LM+SR+1%BC 1±0.2 a 4±0.5 c 20±7 c N.D 91±7 d 136±3

de

LM+SR+5%BC 1±0.3 a 4±0.3 bc 16±1 cd N.D 88±5 d 126±3 e

HM only N.D 20±1 a 113±6 a N.D 198±9 bc 387±13

a

HM+SR 2±0.9 a 5±0.1 b 39±3 b N.D 257±17 a 364±29

a

HM+SR+1%BC 2±0.4 a 4±0.6 bc 20±2 b N.D 251±8 ab 331±8 b

HM+SR+5%BC 1±0.6 a 5±0.2 b 42±4 b N.D 246±13ab 300±12

c

The values are the means of three replicate analyses with “±” indicating standard deviation.

The LSD tests to show the differences between treatments for the same parameter. For the

same fraction across treatments, different letters indicate their significant differences at

P<0.05. “N.D” indicates the “Element cannot be detected under detected limitation”.

45


(magnetite content and SR) on the distribution of As in different chemical forms in LM and

HM tailings (amended and unamended tailings)

Source of

variance

Non-specifically

sorbed

Specifically

sorbed

Sorbed by

amorphous Fe &

Al oxides

Residual

Magnetite 0.043 330.1*** 13316.1*** 51537.5***

SR 10.351** 157.5*** 4151.1*** 989.4*

Magnetite*SR 0.043 276.9*** 5280.8*** 5925.1***

Error 0.554 0.092 11.174 130.6

The values are mean squares. Magnetite in the LM and HM tailings was simply considered as

a treatment factor in the ANOVA (Significant levels: *P< 0.05; ** P< 0.01; *** P< 0.001).


(magnetite content and BC) on the distribution of As in different chemical forms in LM and

HM tailings (only amended tailings)

Source of

variance

Non-specifically

sorbed

Specifically

sorbed

Sorbed by

amorphous Fe &

Al oxides

Residual

Magnetite 0.007 2.407** 3171.7*** 138867.3***

BC 0.226 0.230 16.3 596.4

Magnetite*BC 0.058 0.093 24.862 534.4

Error 0.425 0.167 14.095 577.9

The values are mean squares. Magnetite in the LM and HM tailings was simply considered as

a treatment factor in the ANOVA (Significant levels: *P< 0.05; ** P< 0.01; *** P< 0.001).

46

Table 3-7: Correlation between As fractions in the tailings and tailings properties. The data

from the amended and unamended LM and HM tailings were pooled together in correlation

analysis

Properties Non-specifically

sorbed

Specifically

sorbed

Sorbed by

amorphous Fe &

Al oxides

Residual

pH 0.194 0.313 0.394* 0.375*

EC 0.604** -0.365* -0.349 -0.019

Fe-d 0.299 0.398 0.597** 0.839**

Fe-ox 0.418* -0.219 0.009 0.563**

Mn-d 0.404* -0.190 0.018 0.682**

Mn-ox -0.142 0.852** 0.861** 0.326

Al-d 0.060 -0.209 -0.195 0.085

Al-ox 0.528** 0.179 -0.082 -0.632**

The values are correlation coefficients labelled with their levels of significance (Significant

levels: *P< 0.05; ** P< 0.01; *** P< 0.001).

Fe-d; Mn-d; Al-d: Dithionite extraction-crystalline Fe Al Mn

Fe-ox; Mn-ox; Al-ox: Oxalate extraction-amorphous Fe Al Mn

47

Table 3-8: Correlation among various chemical forms of As in the tailings

Tailing As

fractionation

Non-specifically

sorbed

Specifically

sorbed

Sorbed by

amorphous Fe &

Al oxides

Residual

Non-specifically

sorbed

1

Specifically

sorbed

-0.401* 1

Sorbed by

amorphous Fe &

Al oxides

-0.345 0.942** 1

Residual phase 0.099 0.232 0.502** 1

The data were pooled from the amended and unamended LM and HM tailings for correlation

analysis (Significant levels: *P< 0.05; ** P< 0.01; *** P< 0.001).

3.3 Discussion

3.3.1 Relationship between property changes induced by magnetite recovery and As (V)

adsorption in the tailings

The magnetite recovery process altered the distribution patterns of particle size and decreased

total concentrations of many major elements including As, Fe, Al and Mn in the LM tailings,

but without changing pH and EC conditions in the tailings prior to amendments and

revegetation. The size effects of As-sorbing mineral particles on As adsorption and

desorption in the LM and HM tailings is probably small, although no detailed comparison has

been made. For example, Yean and Cong (2005) found that As adsorption capacity increased

with decreasing magnetite particle size from 20 nm-300 nm, but it did not change greatly

with increasing particle size from µm or mm scales. Mondal and Majumder (2008) found that

As adsorption capacity changed less than 2% when the iron absorbent particle size ranged

from 0.125 mm to 5 mm. The majority (> 95%) of the particles in both the LM and HM

48

tailing samples were not at the nanometre size scale and there is no significant difference in

the clay particles level (0-2 µm) between the two types of tailings.

The magnetite recovery process coincidently decreased total As concentration in the LM

tailings, as indicated by the results. In the magnetite recovery process, the main As-bearing

minerals in the Cu-tailings, arsenopyrite may have been at least partially removed together

with magnetite or leached out in the process. Arsenopyrite occurred at the surface of the

pyrite and had similar stability characters to pyrite with an unstable structure (Craw et al.

2003; Hernández and Canadell 2008). The magnetite contributes to the crystalline Fe fraction.

As a result, the As associated with the crystalline Fe would also have decreased in the process

of the magnetite recovery, resulting in the LM tailings with a lower total As concentration.

Despite this major reduction in total As concentration in the LM tailings, elevated soluble As

levels were observed in the pore water of the LM tailings subject to amendment and


The major changes induced by the magnetite recovery were in the As adsorption capacity and

in the distribution of As into exchangeable and soluble As in the LM tailings, due to the

reduction of As-sorbing Fe-minerals including both crystalline Fe (e.g. magnetite) and

amorphous Fe (e.g. secondary Fe minerals derived from the dissolution of magnetite-Fe),

when the tailings were amended with SR and BC and revegetated with native grass. An

interesting alteration in the distribution of Fe forms occurred in the tailings under direct

revegetation treatments, resulting in increased extractable amorphous Fe, compared to the

unamended treatment in both the LM and HM tailings. These changes may have been caused

by redox changes at magnetite surfaces under well-watered conditions and microbe-mediated

activities which were stimulated by the increased availability of labile carbon in the SR. The

Fe (III) in magnetite surface could be reduced to Fe (II) by dissimilatory iron reducing

bacteria (DIRB) (Dong et al. 2000). Kostka (1995) found that Fe (III) in the magnetite

surface could act as an electron acceptor for the carbon metabolism by reducing bacteria

Shewanella putrefaciens, releasing resultant Fe (II) into the solution phase. In present study,

the Fe (III) exposed at the surfaces of magnetite could have been reduced by iron reducing

bacteria, into the Fe (II), which was precipitated back onto the magnetite particle surface,

forming amorphous Fe-oxyhydroxides. As a result, the amorphous Fe in the HM tailings was

significantly elevated, compared to the LM tailings after the amendment and direct


49

The major differences in As-sorbing Fe-minerals may be the primary reason for the different

As-adsorption capacity between the LM and HM tailings, with much higher arsenate

adsorption capacity in the HM tailings. Even though the amorphous Fe concentrations in the

two tailings were similar, As (V) adsorption could occur on the surfaces of magnetite

particles, even though the affinity of magnetite for As is lower than those of the amorphous

iron minerals (Giménez et al. 2007; Mamindy-Pajany et al. 2011). These changes in As-

sorbing Fe mineral forms and total adsorption capacity have bearings on the distribution and

solubility of As in the Cu-tailings after the recovery of magnetite.

3.3.2 Arsenic distribution and re-distribution in the tailings

The distribution of As forms in the tailings was significantly altered by the factors of

magnetite content and organic matter amendments under direct revegetation. The changes in

As form distribution in the tailings seem to be coupled with the changes of Fe-forms, which

is closely related to the activity of Fe-reducing and oxidising bacteria, since Fe minerals are

the major adsorbing phase in the tailings. With the dissolution of Fe minerals (such as

magnetite, hematite, etc.) in the tailings catalysed by Fe-reducing bacteria, As sorbed to the

Fe minerals can be released into the pore water due to the loss of the Fe-bonding phase (Dong

et al. 2000; Yamaguchi et al. 2011). In neutral or near neutral pH conditions, Fe minerals are

seldom soluble under normal redox conditions and therefore As sorbed by the Fe minerals is

barely soluble in the tailings (Straub et al. 2001). This is similar to previous findings that soil

and tailings may contain high concentrations of total As but little (NH4)2SO4 extractable As

(<1 mg kg-1

air dwt and < 1% (less than) of the total As) (Ascar et al. 2008a; Krysiak and

Karczewska 2007; Lee et al. 2010). The highest extractable As fraction (except for the

residue As) was associated with the amorphous Fe Al Mn minerals, which highlighted their

important role in As immobilisation in the tailings. This phenomenon is consistent with

previous findings (Bowell 1994; Krysiak and Karczewska 2007; Matera et al. 2003; Voigt et

al. 1996).

The strong As adsorption capacity in the HM tailings may be due to the formation of As-

binding amorphous Fe at the surfaces of magnetite particles. The iron reducing bacteria

Shewanella putrefaciens could metabolise the Fe (III) on the surface of magnetite and form a

layer of ferrihydrides and FeCO3 coating on magnetite surface which has very high affinity

for As in solution (Roden and Zachara 1996). This As may be associated with the crystalline

silicate lattices and barely be extracted by chemical methods (Samanidou and Fytianos 1987).

50

The As distribution in the solid phase of tailings is related to the tailing properties, mainly the

forms and quantity of As-sorbing minerals such as magnetite and its derivatives after redox

changes at its particle surfaces. Among these properties, the As distribution in Fe phases was

significantly correlated with the crystalline Fe in the tailings. In each As fractionation phase,

the specifically sorbed As was highly and positively correlated to the amorphous iron phase

(r=0.942).

Amorphous Mn has also contributed to the As distribution in the specifically sorbed and

amorphous Fe phase in the tailings. The amorphous Mn in HM tailings was approximately

three times more than that in the LM tailings. Amorphous Mn oxides also have a significant

As-adsorption capacity (Foster et al. 2003; Scott and Morgan 1995). Although amorphous

hydrous Mn oxides exhibit a lower As (V) adsorption capacity in (HMO) than amorphous Fe

oxides, studies have found that they remained relatively effective in As adsorption under

certain conditions (Balistrieri and Chao 1990; Foster et al. 2003). Takamatsu (1985) found

that more than 80% of As (V) could be absorbed to hydrous Mn oxides at pH 4.

In addition, phosphate (PO43-

or H2PO4-) competes with As (V) for sorption sites at the

surfaces of Fe minerals (Jain and Loeppert 2000; Manning and Goldberg 1996), which should

be considered when adding P-fertilisers in the amended tailings for plant growth. The

increased specifically sorbed As in the LM tailings with OM treatments indicated that the As

could be more likely to be exchanged by phosphate and leached into the pore water compared

to the HM tailings, as iron oxides have higher affinity for phosphate than arsenate. In the

residue phase, As was in a higher proportion than other extraction phases. This specific

relationship should be investigated further in the amended LM tailings under direct

revegetation.

3.4 Summary

The magnetite recovery practice has changed the properties of the Cu-tailings at EHM, which

substantially decreased the total element (Fe, As, Al, Mn) concentrations in the LM tailings,

compared to that in HM tailings. The altered composition of mineralogy affected As

mobilisation and distribution in the tailings. The HM tailings had a significantly higher As (V)

adsorption capacity than the LM tailings. In LM tailings, with the extraction of the crystalline

Fe oxides (magnetite), most of the Fe-bonding As was associated with amorphous Fe oxides.

The As distribution in different chemical forms in the tailings can be attributed to the iron

transformation catalysed by reducing and oxidising bacteria. The reduced As (V) absorption

51

capacity in the LM tailings may favour more As distribution into the pore water and thus

increase As availability for plant uptake when the LM tailings are under amendment and

direct revegetation. This hypothesis has been investigated in Chapter 4.

52

Chapter 4 Arsenic dissolution and speciation in pore water of high and low magnetite

tailings amended with organic matter

4.0 Introduction

The presence of soluble inorganic As species in the pore water of amended tailings is the

prerequisite for plant uptake, as the preferred forms for plant roots are the inorganic As

species, which form via speciation (including inorganic or organic As species) in the

rhizosphere (Zhao et al. 2010). The uptake of arsenate (As (V)) in roots shares the same

cellular pathway with phosphate, while plant roots adsorb As (III) in the form of neutral

(H3AsO3) molecules (Asher and Reay 1979; Chen et al. 2005). The soluble inorganic As in

the pore water is the result of the dissolution of As-bearing minerals (such as arsenopyrite in

Cu mine tailings) in the oxidation process and accompanying adsorption by secondary

Al/Fe/Mn oxides and oxyhydroxides formed through the transformation of Fe-minerals.

These may then be further converted into organic As through microbial-mediated

transformation processes in the amended tailings (Al-Abed et al. 2007; Drahota and Filippi

2009; Zhao et al. 2010). As a result, As speciation in the rhizosphere solution is closely

regulated by pH and redox conditions, the presence of Fe oxyhydroxides and the activities of

oxidising and reducing bacteria involved in the redox processes of Fe and As. In particular,

inorganic As concentration in pore water is closely dependent on the concentration of

adsorbing sites present on the surfaces of Fe-minerals (including amorphous and crystalline

forms) in tailings and soils (Drahota and Filippi 2009).

Therefore, contents of primary Fe-minerals (such as magnetite) in the tailings would have

significant bearing on As adsorption capacity and the concentration of soluble As in the pore

water for plant uptake. This is because primary Fe minerals such as magnetite can be

modified in microbe-mediated processes to form secondary Fe minerals (including Fe

oxyhydroxides) which precipitate and coat the surfaces of the primary Fe mineral particles.

As described in Chapter 3, the distribution of As in the soluble pool greatly increased in the

low magnetite (LM) Cu-tailings compared to the high magnetite (HM) tailings. As a result,

we expect that plants grown in the LM tailings would take up more As in the biomass than

those in the HM tailings.

In addition to the total soluble As concentration, the varying chemical forms or species of As

in the pore water are also closely related to plant uptake, which itself is closely influenced by

53

the activity of microbes involved in As methylation and formation of organic As (Zhao et al.

2010). By adding organic matter to the tailings, the activities of reducing and oxidising

bacteria can be increased by labile carbon resources (C6H12O6). This can then stimulate the

dissolution of Fe from the surfaces of magnetite particles to form secondary Fe minerals

(such as Fe oxyhydroxides) and the co-dissolution of As from arsenopyrites and/or As-Fe-

secondary minerals (Morin and Calas 2006; Powlson et al. 1987). For example, Fe and

associated heavy metals such as Cd, Cr, Ni, Pb in goethite may be released into the solution

phase due to the dissolution of Fe-oxides, which are catalysed by Fe-reducing bacteria

(Francis and Dodge 1990). In sulphidic tailings, Fe-oxidising bacteria such as Leptospirillum

ferrooxidans and Ferroplasma acidiphilum are dominant in microbial communities; these

bacteria could directly stimulate the pyrite oxidation (Johnson et al. 2001; Mendez et al.

2008). The end products of asenopyrite oxidation involve ferric, arsenate, and sulphate as

well as H+ (Corkhill et al. 2008).

Organic matter amendment is a common practice to improve mine tailings for vegetation

establishment. Many different types of organic matter have been used to improve the physical

and chemical conditions in the tailings to increase the survival rate of plants, including

woodchips, biosolids and biochars (Fellet et al. 2011; Hulshof et al. 2006; Van Rensburg and

Morgenthal 2004). Different organic matters have different proportions of labile and stable

carbons, for example sugarcane residue contains large amounts of labile carbon, while

biochar consists of mainly stable carbon. The addition of different organic matter may induce

different hydro-geochemical reactions and the evolution of different microbial communities

and activities (Hulshof et al. 2006; Li et al. 2013).

The increased organic labile carbon in the tailings may stimulate the activity of Fe-

reducing/oxidising bacteria and thus increase Fe dissolution or transformation into other Fe-

oxyhydroxides, which are strong sinks for inorganic As. It may also stimulate activities of

microbes involved in the redox changes of inorganic As and As methylation. Arsenic species

in the pore water are the consequence of microbial-mediated reduction, oxidation and

methylation processes (Stolz and Oremland 1999; Zhao et al. 2010). From this we can expect

that organic matter amendments may stimulate the dissolution of Fe at the surface of

crystalline Fe minerals such as magnetite and hematite. This may then be precipitated as Fe-

oxyhydroxides, which coat the surfaces and adsorb As from the solution. As a result, there

would be more inorganic As distributed in the pore water of the LM tailings due to the lower

54

As adsorption capacity, compared to the HM tailings. Consequently, the conversion of

inorganic As into organic As species may be enhanced in the LM tailings, when they are

amended with organic matter with high labile carbon sources. Organic matter may also

change redox conditions in the tailings and the dissolution of Fe-minerals and thus alter As

solubility and speciation in the pore water (Redman et al. 2002; Zachara et al. 2002).

Meanwhile, the associated As dissolution with the Fe-mineral dissolution could not be ruled

out due to the co-occurrence of As and Fe in these minerals in the Cu tailings.

The present chapter aimed to investigate effects of combined amendments with sugarcane

residue (SR) and different rates of biochar (BC) on As solubility and speciation in the pore

water and the resultant As uptake by a native grass (Red Flinders grass) grown in the two Cu

tailings (LM and HM) under glasshouse conditions. Preliminary experiments showed that the

plants could not survive without the addition of sugarcane residue. As a result, the addition of

SR was applied as a basal treatment in the plant trial, with the rate of BC and tailings type as

treatment factors. In detail, the present study investigated the effects of combined treatments

of SR and BC treatments in the LM and HM tailings, on (1) the soluble As and Fe

concentrations in pore water; (2) the distribution of As among inorganic and organic species

in the soluble phase; (3) As uptake and distribution in native grass (Red Flinders grass) under

direct revegetation.

4.1 Materials and Methods

4.1.1 Plant culture and treatment

Detailed information can be found in Chapter 3, regarding the collection of tailings and the

basic properties of the two types of Cu-tailings. The LM and HM tailings were amended with

two types of organic matter including sugarcane residue (SR) (Earth Wise Company, Qld)

and pine biochar (BC) (ANZAC Pty Ltd, pyrolysed at 700 °C), which were dried at 60 °C for

3 days before use. The organic matters were ground and sieved through 2 mm sieves to

produce consistent mixtures in the amendment treatments.

Organic matter amendments were applied on a weight basis (% w/w) in the LM and HM

tailings by thoroughly mixing appropriate amounts of SR and/or BC into 3 kg lots of the

tailings in pots. The treatments were: unamended (control), SR 5% only, SR 5%+BC 1%, and

SR 5%+BC 5%. The control treatments were only replicated in two pots, while the organic

matter amendment treatments were replicated in four pots.

55

The glasshouse experiment was conducted between February to May 2013 with an ambient

temperature of 26-32°C. Seeds of the native Red Flinders grass (Iseclema Vagin florum) were

supplied by the Ernest Henry Mine, which were further selected from the mature flowers

manually. The seeds were directly germinated in potting mix soil (sand and soil mixture 1:3

which was steam sterilised) and watered daily. Germinated seedlings were cultured under

glasshouse conditions prior to transfer into the treatments. At Day 20 after germination,

seedlings of similar size and appearance were transplanted into the pots containing the

treatments as described above, at the rate of six seedlings per pot that were randomly selected

for consistency. The plants were grown for 4 weeks after transplanting. At transfer, seedlings

were gently excavated from the seeding tray to minimise root damage and their roots were

gently washed with deionised water to minimise the carryover of potting mix soil into the

tailings treatments. The pots were laid out on bench randomly and rotated each week.

The plants were grown in a twin-pot-system with capillary watering mechanism (Adam et al.

2014). The system contained a top pot (20 cm high, 15 cm diameter) with the tailings

treatments and bottom pot (15 cm high, 12cm diameter) containing water (see Chapter 6

twin-pot diagram). The pots were connected with a capillary mat for water supply into the

root zone via a capillary mechanism. Approximately 1.5 L 50% (diluted with deionised water)

fill-strength Hoagland solution (see Table 4-1 for nutrient composition) was supplied in the

bottom bucket, which was refilled weekly.

Table 4-1: The nutrient solution used to irrigate plants in the glasshouse “twin-pot” system

Macronutrients (mM) Micronutrients (µM)

KNO3 0.25 FeEDTA 2

Ca(NO3)2. 4H2O 0.125 ZnSO4.7H2O 0.1

MgSO4.7H2O 0.05 MnSO4.H2O 0.1

KH2PO4 0.005 CuSO4.5H2O 0.025

K2HPO4 0.005 Na2MoO4.2H2O 0.004

NaCl 0.4

H3BO3 0.262

56

4.1.2 Pore water sampling and chemical analysis

For sampling and monitoring chemical changes in the pore water of the tailings treatments

during the experimental period, pore water samplers (Rhizon MOM, 10 cm porous, glass

fiber, Rhizosphere Research Products, Wageningen, Netherlands) were inserted into the

middle layer of the tailings profile (at about 7 cm from the bottom of a pot). This allowed the

continuous collection of pore water samples without affecting plant roots (Xu et al. 2008).

The holes into which the samplers were inserted on the side of each pot were sealed with

silicon glue to hold the samplers in position. Pore water collection usually started at 7 am on

the day, due to low evaporation rate at this time. A 5 ml syringe was connected to each

sampler top and its plunger was pulled outwards to a fixed volume mark and fixed with a

stopper to create a vacuum pressure (less than 50 kPa) (Wopereis 1994). The pots were

sampled for pore water once per week for a four week period. The pore water samples from

each pot filled in the syringe under the gentle vacuum (less than 5 minutes) and were

transferred into 10 ml polystyrene vials with an airtight cap. Pore water samples of each pot

at each time were sampled in duplicates and immediately transported back to the laboratory

in a double walled plastic container precooled to approximately 5-8 °C. Aliquots of one of

the duplicate pore water samples were used for pH and electrical conductivity (25 °C)

measurements. In supplementary pots, additional pore water samples were collected from the

pots for immediate redox potential measurements in the pore water samples. The redox

potentials in this additional set of pore water samples were measured by quickly inserting the

ORP (redox) sensor immediately after collection without agitation. This was a compromise

due to the unavailability of in situ sensors at the time, which could be placed in the tailings

profile for accurate measurements over time.

From the duplicate pore water samples of each pot, one set was immediately acidified with

concentrated nitric acid (0.08 ml per 10 ml sample) and stored in a cold room (4 °C) for total

elemental (including As) analysis using inductively coupled plasma – optical emission

spectroscopy (ICP-OES, Perkin Elmer) or ICP-Mass Spectroscopy (Agilent Technologies),

depending on elemental concentrations and detection limits. The other set of the same pore

water sample of each pot was stored in a freezer (-18 °C) immediately after collection for As

speciation using the HPLC-ICP-MS method (Agilent Technologies) (Van den Broeck et al.

1998). The samples were acidified with 0.08 ml concentrated HCl/10 ml sample solution

immediately after thawing to prevent the co-precipitation of As with the formed Fe-

oxyhydroxides, as the pore water samples contain high levels of soluble Fe under the

57

watering conditions in the glasshouse (based on our preliminary tests). The As species (As

(III), As (V), MMA, DMA, AsB) were separated with an anion-exchange As speciation

column and fitted with a guard column. The mobile phase was the solution prepared with

2mM NaH2PO4 and 0.2 mM Na2-EDTA, which was pumped through the column at 1ml per

minute (Xu et al. 2007). The separated column solution was connected with the ICP-MS for

As analysis.

4.1.3 Plant harvest and analysis

Plant shoots (1-2 seedling, including dried leaves) were thinned in the second and third week

to minimise inter-plant competition in each pot. The remaining plants (2-3 per pot) were

destructively harvested in the fourth week. The plants thinned at each week were identified

on the basis of the physiological growth status. The shoots were cut at the base (near the

surface of the tailings), which were thoroughly rinsed with deionised water to prevent

contamination of the plant samples by tailings particles. The plant samples were blotted dry

with paper towel and placed in paper bags for drying at 65 °C for until a constant weight was

reached. Roots were only harvested in the fourth week, after the shoots had been harvested in

the manner as described. The roots were slowly separated from the tailings by gently washing

under running tap water. After initial separation, the roots were further washed with

deionised water for more than 10 changes of DI water, until the absence of tailings particles

based on visual observations. The washed roots were blotted dry with paper towel and placed

in the paper bag for drying at 65°C until a constant weight was reached.

The dried shoots and roots biomass were accurately weighted (0.1g for shoots and 0.05g for

roots) to determine biomass. The whole shoot and root samples were digested in concentrated

nitric acid without grinding (as the samples were small) using a microwave open vessel

digestion method by means of a Milestone microwave digestion system (Huang et al. 2004;

Lamble and Hill 1998). After dilution with Millipore water, the digested sample solution was

analysed for total Fe and As concentrations using ICP-MS. At least three blank samples and

the reference standard plant sample Apple Leaves (NIST SRM 1515) were included in each

batch for quality control purposes.

4.1.4 Data analysis

Two-way analysis of variance was carried out to evaluate treatment effects and their

interactions (SPSS 20.0, IBM, USA). The differences among the means were compared using

LSD-0.05. Where appropriate, the data were LOG10 transformed before carrying out ANOVA

58

and LSD analysis. Correlation analysis was performed to evaluate the relationships among

various sets of parameters as indicated in relevant tables and figures.

4.2 Results

4.2.1 Pore water properties in EHM tailings

4.2.1.1 Changes of pH, EC and Redox potential

The pore water pH conditions in both LM and HM tailings were neutral or near neutral from

first week to fourth week (Figure 4-1). Pore water pH conditions showed a slightly

decreasing trend in the LM tailings treatments, but they remained stable in the HM. In the BC

treatments, 1% BC in both LM and HM tailings showed higher pH at 7.02 and 7.06 in first

week. In the fourth week, pH in 1% and 5% BC treatments was 0.5 higher than in the 0% BC

treatment in LM tailings. No obvious differences between treatments in pore water pH were

observed in the HM tailings. In the control treatments without SR and BC treatments, the

pore water pH remained stable, which was around 7.35 and 7.45 in the LM and HM tailings,

respectively. The pore water EC (ms cm-1

) in LM tailings decreased with the increasing BC

levels (Figure 4-2). In detail, the EC levels decreased from 4.08 to 3.53, 3.50 to 3.09 and

2.23 to 2.61, in 0% BC, 1% BC and 5% BC treatments, respectively. In the pore water of the

HM tailings, the EC trend did not reflect the trend in the LM tailings. The 0% BC treatment

in the HM tailings had the highest EC level. In the control treatments, the pore water EC

values (ms cm-1

) remained stable in this period, which were 4.11 in LM tailings and 3.41 in

HM tailings. The redox status in the pore water remained stable during the glasshouse

incubation period (Figure 4-3). Pore water redox results both ranged from 185 to 204 mv in

the organic matter amendment treatments in two tailings, while it was around 178 to 193 mv

in the control treatments without any amendments in both the LM and HM tailings.

59

Figure 4-1: Changes of pH conditions in the pore water of the amended LM and HM tailings

during the 4 weeks of glasshouse incubation

Figure 4-2: Pore water EC in the LM and HM tailings during the period of glasshouse

experiment

6.4

6.7

7

7.3

1st week 2nd week 3rd week 4th week

pH

(unit

)

Time after commencing treatment

LM+SR LM+SR+1%BC LM+SR+5%BC

HM+SR HM+SR+1%BC HM+SR+5%BC

2.5

3

3.5

4

4.5

1st week 2nd week 3rd week 4th week

EC

(m

s cm

-1)

Time after commencing treatment

LM+SR LM+SR+1%BC LM+SR+5%BC

HM+SR HM+SR+1%BC HM+SR+5%BC

60

Figure 4-3: Redox potentials in the amended tailings under well watered conditions in the

glasshouse experiment, which were measured in freshly collected leachate of 1 pot twice a

week after commencing treatment

4.2.1.2 Soluble As and Fe concentrations in pore water

Total As concentrations in the pore water from the LM tailings were significantly higher than

those in the HM tailings, which increased with the treatment time in both tailings. In the first

and second week, the total pore water As concentrations were less than 50 µg L-1

in the LM

and HM tailings amended with organic matter (except for LM+SR treatment in second week)

(Table 4-2). In the third and fourth week, total As concentrations in the pore water increased

dramatically, compared to those in week 1 and 2. In fourth week, total As concentration in

the pore water of the LM tailings was about 6-7 fold higher than those of the HM tailings.

The average As concentrations were 688 µg L-1

, 644 µg L-1

, 406 µg L-1

in the LM tailings

amended with SR and 1% or 5 % BC, respectively; while they were 34 µg L-1

, 62 µg L-1

and

60 µg L-1

in the HM tailings. There was no significant difference in the various different BC

treatments. In the control treatments, total pore water As concentration remained extremely

low across the 4-week period, which was as low as 2.3 µg L-1

and 1.6 µg L-1

in LM and HM

tailings, respectively.

The trend of total Fe concentrations in the pore water samples was similar to that of As

concentration changes in both tailings, which increased with time. Total levels of soluble Fe

in the pore water of LM tailings were consistently and significantly higher than those of the

0

50

100

150

200

250

300

SR SR+1%BC SR+5%BC

Red

ox

(m

v)

Treatment

LM tailings

HM tailngs

61

HM tailings throughout the experiment (Table 4-3). By the third to fourth week, the Fe

concentration in the pore water samples of the LM tailings rose to about 270 mg L-1

compared to 35-60 mg L-1

in the HM tailings. In the control tailings without any amendments

and plants, Fe concentrations in the pore water samples were merely 0.04 and 0.05 mg L-1

in

LM and HM tailings, respectively.

By comparing the main effects of treatment factors and their interactions, total As and Fe

concentrations of the pore water in the LM and HM tailings were significantly (p<0.001)

affected by the levels of magnetite in the tailings, while BC addition had little effect. There

were no interactions between the two main factors (magnetite content and BC) (Table 4-4).

Table 4-2: Total As concentration (µg L-1


with organic matter

Time 1st week 2

nd week 3

rd week 4

th week

Treatments µg L-1

LM+SR 13±1 ab 129±36 a 505±11 a 688±44 a

LM+SR+1%BC 12±2 ab 17±1 b 538±302 a 711±425 a

LM+SR+5%BC 14±6 ab 28±11 b 516±196 a 406±138 a

HM+SR 20±2 a 21±3 b 34±10 b 34±10 b

HM+SR+1%BC 18±6 a 19±2 b 27±4 b 62±32 b

HM+SR+5%BC 10±1 b 11±1 c 54 ±24 b 60±32 b

The values are the means of three replicates ± standard deviation. The LSD tests were used to

compare the differences between treatments for the same period of treatment time (e.g. 1st,

2nd

, 3rd

or 4th

week after commencing treatments). For the same period across treatments in

each column, different letters indicate their significant differences at P<0.05.

62

Table 4-3: Total Fe concentration (mg L-1


with organic matter

Time 1st week 2

nd week 3

rd week 4

th week

Treatments mg L-1

LM+SR 17±4 a 141±30 a 245±17 a 270 ±13 a

LM+SR+1%BC 0.3±0.1 cd 1.2±0.7 bc 166±66 a 182±118 a

LM+SR+5%BC 3±0.7 b 51±32 ab 181±51 a 154±48 a

HM+SR 0.1±0.0 de 5±1 bc 31±7 b 31±13 ab

HM+SR+1%BC 0.2±0.2 d 9±6 bc 27±8 b 36±33 b

HM+SR+5%BC 0.9±0.4 c 6±2 b 37±22 b 43±40 ab


compare the differences between treatments for the same period of treatment time (e.g. 1st,

2nd

, 3rd

or 4th

week after commencing treatments). For the same period across treatments in

each column, different letters indicate their significant differences at P<0.05.


(magnetite content and BC) on total As and Fe concentrations in pore water of the LM and

HM tailings

Source of variance Total As Total Fe

Magnetite 7.4*** 14.1***

BC 0.11 3.2

Magnetite*BC 0.11 1.2

Error 0.32 0.96

The values are mean squares. Magnetite content in the LM and HM tailings was simply

considered as a treatment factor in the ANOVA (Significant levels: *P< 0.05; ** P< 0.01;

*** P< 0.001).

63

4.2.1.3 Arsenic speciation in pore water

The pattern of As species distribution in the pore water samples appeared to differ between

the LM and HM tailings in response to the amendment treatments and plant growth. The

concentrations of inorganic (As (III) and As (V)) and organic As species (DMA) in LM

tailings were significantly elevated compared to those in the HM tailings subject to the same

amendment and plant treatments (Table 4-5). In the pore water samples of LM tailings

subject to the organic matter treatments, As (III) and As (V) concentrations measured 220-

400 µg L-1

and As (V) 105-220 µg L-1

. In contrast, As (III) and As (V) concentrations in the

HM tailings amended with organic matter were much lower, measuring 10-26 and 12-30 µg

L-1

, respectively, even though the HM tailings contained more total As than the LM. Overall,

the levels of organic As species were lower than those of the inorganic in both tailings. The

concentrations of AsB and DMA in the pore water of the LM tailings were 3-7 fold higher

than those of the HM.

The removal of magnetite from the Cu-tailings was the major cause for the elevated

concentrations of As (III), As (V), AsB and DMA in the pore water of LM tailings, compared

to those in the HM (P<0.001 for As (III), As (V) and DMA and P<0.01 for AsB). The effects

of BC amendment and interactions between the magnetite content and BC were not

significant (Table 4-6).

The relative distribution patterns of As species in the soluble As pool should be interpreted

against their actual concentrations in the pore water samples as described above (Figure 4-4).

The proportion of As (III) in the pore water of the LM tailings was higher than those of the

HM in all the treatments, namely 49%, 28% and 20% higher in SR, SR+1%BC and

SR+5%BC, respectively. The inorganic As species including As (III) and As (V) accounted

for the major proportions of the total soluble As in the pore water of the LM tailings, which

measured 94%, 80% and 83% in SR, SR+1% BC, SR+5% BC treatments, respectively. In the

pore water of HM tailings, there was an increasing trend of As distribution into the inorganic

forms from 44% to 77%.

64

Table 4-5: Concentrations of As species in the pore water collected in the 4th

week from the

LM and HM tailings amended with organic matter

As speciation As (III) As (V) AsB DMA

Treatments µg L-1

LM+SR 402±57 a 225±8 a 19±10 a 22±5 b

LM+SR+1%BC 406± 298 a 202±140 a 32±7 ab 82±35 a

LM+SR+5%BC 223±53 a 104±13 a 36±43 ab 41±26 b

HM+SR 18±14 b 12±6 b 10±3 b 10±4 c

HM+SR+1%BC 18± 6 bc 30±20 b 6±1 bc 8±2 c

HM+SR+5%BC 26±20 c 27±16 b 6±1 bc 3±1 d

The values are the means of three replicates ± standard deviation. The LSD tests were

conducted to compare the differences between treatments for the same As species. For the

same As species across treatments in each column, different letters indicate their significant

differences at P<0.05.


(magnetite content and BC) on the distribution of soluble As among different As species in

the LM and HM tailings amended with organic matter

Source of

variance

As (III) As (V) AsB DMA

Magnetite 10.5*** 4.1*** 1.2** 2.9***

BC 0.11 0.03 0.01 0.22

Magnetite*BC 0.37 0.15 0.08 0.21

Error 0.13 0.08 0.06 0.03

The values are the mean squares. Magnetite in the LM and HM tailings was simply


*** P< 0.001).

65

Figure 4-4: The distribution of soluble As among different As species in the pore water of the

LM and HM tailings amended with combinations of sugarcane SR and BC, (LM refers to low

magnetite tailings, HM refers to high magnetite tailings)

0%

20%

40%

60%

80%

100%

SR SR+1%BC SR+5%BC

Treatment

(A) LM

DMA

AsB

As (V)

As (III)

0%

20%

40%

60%

80%

100%

SR SR+1%BC SR+5%BC

Treatment

(B) HM

66

4.2.2 Plants response to tailings

4.2.2.1 Plant Biomass

Plant biomass did not show a significant response to the organic matter (SR and BC)

treatments in both LM and HM tailings (Table 4-7). In general, there was no significant

effect of the treatments and tailings types on plant shoot and root biomass except the

treatment of HM+SR+1%BC. There was no interaction between the two main factors

(magnetite content and BC) on plant growth (Table 4-8).

Table 4-7: The biomass of Red Flinders grass grown in organic matter amended LM and HM

tailings for 4 weeks

Treatments Root (g/dry wt) Shoot (g/dry wt) Total (g/dry wt)

LM+SR 0.37±0.09 a 1.39±0.26 b 1.76±0.34 b

LM+SR+1% BC 0.32±0.09 a 1.84±0.46 b 2.17±0.54 ab

LM+SR+5% BC 0.32±0.06 a 1.51±0.27 b 1.82±0.32 b

HM+SR 0.27±0.03 a 1.61±0.38 b 1.87±0.40 b

HM+SR+1% BC 0.31±0.09 a 2.54±0.47 a 2.90±0.58 a

HM+SR+5% BC 0.26±0.02 a 1.48±0.19 b 1.74±0.21 b


compare the differences between treatments. For the same plant part across treatments,

different letters indicate their significant differences at P<0.05.

67


(magnetite content and BC) on the plant biomass of Red Flinders grass grown in the amended

LM and HM tailings

Source of variance Root Shoot Total biomass

Magnetite 0.015 0.104 0.026

BC 0.001 0.593 0.621

Magnetite*BC 0.009 0.281 0.357

Error 0.007 0.189 0.256


considered as a treatment factor in the ANOVA. No significant effects were produced by the

treatments under glasshouse conditions (Significant levels: *P< 0.05; ** P< 0.01; *** P<

0.001).

4.2.2.2 Plant As uptake

The Red Flinders grass in the amended LM tailings absorbed more As in roots and shoots

than those in the HM (Figure 4-5). Total As concentration in the shoots ranged from 0.7 to

2.2 mg kg-1

dwt in the organic matter treatments of LM and HM tailings. In the 1% BC

treatment, the grass grown in LM tailings accumulated more As in the shoots compared to

those in the HM tailings. Arsenic concentrations in roots were much higher than those in

shoots, which was consistent between the LM and HM tailings, regardless of the organic

matter treatments. Root As concentration in the LM tailings was 8-10 fold higher than those

in the HM tailings across the treatments. The highest As concentration in roots was 146 mg

kg-1

dwt in the LM tailings with 1% BC, while the lowest As concentration was 9 mg kg-1

dwt in in the HM tailings with SR only. Arsenic concentration in the roots grown in the LM

tailings with 5% BC was significantly lower than those amended with SR and 1% BC

treatment. However, this trend was not apparent in the HM tailings.

By comparing the main effects of treatment factors and their interactions, the As uptake by

shoots and roots of Red Flinders grass grown in the amended tailings was significantly

(p<0.001 for roots and p<0.01 for shoots) affected by the removal of magnetite from the Cu-

tailings. The effect of BC amendment on root As concentration was significant (p<0.05).

There were no interactions between the two main factors (magnetite content and BC) (Table

68

4-9). From the correlation analysis, we found that the As accumulation in shoot of Red

Flinders grass grown in LM and HM organic matter amended tailings was significantly

(p<0.05) and positively related to the pore water concentrations of total As, As (III), As (V)

and total P concentrations in the pore water (Table 4-10). Arsenic accumulation in the roots

was significantly (p<0.01) and positively related to the concentrations of total As, As (III),

As (V), AsB, DMA and total P in the pore water (Table 4-10).

Figure 4-5: Arsenic concentrations in the shoot and root of Red Flinders grass harvested in

the 4th

week after commencing treatments of organic matter amendments in the LM and HM

tailings

a

a

a ab

b ab

0

1

2

3

SR SR+1% BC SR+5% BC

As

upta

ke

(mg k

g-1

dry

bio

mas

s)

Treatment

(A) Shoot

LM tailings

HM tailings

a a

c

b b b

0

30

60

90

120

150


As

upta

ke

(mg k

g-1

dry

bio

mas

s)

Treatment

(B) Root

69


compare the differences between treatments and different letters indicate their significant



(magnetite content and BC) on As concentrations in the shoot and root of Red Flinders grass

grown in the LM and HM tailings amended with biochar (BC). Sugarcane residue was used

as a basal amendment across the treatments

Source of variance Root As Shoot As

Magnetite 34566*** 1.56**

BC 1124* 0.082

Magnetite*BC 809 0.054

Error 278 0.123



*** P< 0.001).

Table 4-10: Correlation between As concentrations in Red Flinders grass and concentrations

of total As and As species in the pore water of the LM and HM tailings amended with organic

matter

Plants As

uptake

Total As As (III) As (V) AsB DMA Pore water

total P

Shoot 0.454* 0.489* 0.446* 0.407 0.467 0.446*

Root 0.735** 0.712** 0.701** 0.524* 0.547** 0.641**

The values are correlation coefficients labelled with their levels of significance (Significant

levels: *P< 0.05; ** P< 0.01; *** P< 0.001).

70

4.3 Discussion

4.3.1 Arsenic and Fe dissolution in pore water of LM and HM tailings amended with

organic matter

The present findings have confirmed the initial hypothesis that low magnetite in the tailings

would favour the dissolution of As minerals and the distribution of soluble As in the total As

pool and the increased organic As formation in the LM tailings. The organic matter

amendment in the Cu-tailings stimulated the dissolution and mobilisation of both As and Fe,

as indicated by the strong positive correlation (P<0.05) between soluble Fe and As

concentrations in the pore water, regardless of the magnetite contents in the tailings (Figure

4-6). Organic matter such as SR can provide labile carbon and enhance the microbe activity

in tailings, which could be utilised by the bacteria as electron acceptors (Mendez and Maier

2008; Robinson et al. 2001). Lovely (1998) found that Fe (III) oxide reduction and thus

dissolution were both stimulated by the Fe-reducing bacteria (Geobater metallireducens)

when exposed to humic substances. In this study, As barely dissolved in the pore water of the

control treatment without any organic matter amendment and plants. Further, Fe-oxidising

bacteria could oxidise Fe2+

to Fe3+

quickly and form ferrihydroxides at the surface of the Fe

minerals. The oxidation progress proceeded slower under abiotic (control) conditions

(Rimstidt and Vaughan 2003; Rojas‐Chapana and Tributsch 2004). It was well known that

iron oxides and hydroxides play a significant role in As retention (Giménez et al. 2007). This

may be the reason that more soluble As was present in the pore water of the LM tailings,

compared to those in the HM.

These changes of Fe dissolution and precipitation may have occurred in the Cu-tailings

amended with organic matter under revegetation in the present study, which resulted in

modified As adsorption capacity and As distribution into the pore water. Crystalline Fe

minerals such as magnetite are thermodynamically more stable than amorphous Fe such as

ferrihydrite and lepidocrocite (Cornell and Schwertmann 2003; Pedersen et al. 2006).

Crystalline Fe minerals have much lower As-bonding ability due to their lower specific

surface area than the amorphous forms (Kocar et al. 2006). Fe located at the surfaces of

magnetite particles could be dissolved by the bacteria Shewanella putrefaciens and release

Fe2+

, which, in the presence of HCO3−

may form siderite (Dong et al. 2000; Kostka and

Nealson 1995). Studies have shown that the siderite and the carbonates form precipitates to

coat the surface of the primary minerals and prevent further dissolution of these minerals (Al

71

et al. 2000). In weathered Cu-tailings, the presence of amorphous gel of ferrihydrites was

observed on pyrite mineral particles (Forsyth 2010). The large surface area from ferrihydrite,

gives rise to a strong adsorption capacity for phosphate (O’Loughlin et al. 2013) and arsenate

due its similar adsorption characteristic to phosphate (Zhao and Stanforth 2001). After

extensive weathering, amorphous Fe oxides such as ferrihydrite may be transformed into

crystalline forms such as hematite, goethite and magnetite, by bacteria, which would result in

reduced specific surface area and decreased As adsorption on the minerals (Hansel et al. 2003;

Pedersen et al. 2005; Tamaura et al. 1983; Zachara et al. 2002).

As a result, the redox conditions and the activities of reducing bacteria in the tailings

amended with organic matter may have generated different effects of Fe dissolution and

precipitation on magnetite particle surfaces in the LM and HM tailings. The presence of high

magnetite contents would be translated into higher surface areas coated by amorphous Fe-

minerals and thus higher As-adsorption capacity in the HM tailings compared to that of the

LM (see Chapter 3). The high As adsorption capacity in the HM tailings resulted in the

significantly lower soluble As concentration in the pore water of the tailings, despite organic

matter amendment and anaerobic conditions.

Nevertheless, magnetite particles also have certain levels of As adsorption capacity, as

demonstrated in wastewater studies (An et al. 2011; Chandra et al. 2010; Ohe et al. 2005;

Yoshizuka et al. 2010). In Chapter 3, it was already suggested that the high magnetite HM

tailings had a stronger As (V) adsorption capacity than that of LM tailings. Hence, the high

level of magnetite itself may act as a sink to prevent As mobilisation in HM tailings. Overall,

the mobilisation of As and Fe were both suppressed by the HM in copper tailings.

72

Figure 4-6: The relationship between total soluble As concentration and soluble Fe

concentration in the organic matter amended LM and HM tailings from the glasshouse

experiment

4.3.2 Arsenic speciation in the pore water and Fe mineral forms

The relative patterns of As speciation and distribution of the soluble As pool were affected by

the differences in magnetite content, because of the As adsorption capacity in the LM and

HM tailings. In the LM tailings, the elevated levels of soluble As in the pore water provided

substrate for bacteria-mediated As transformation from inorganic into organic As forms. In

the present study, the amount of inorganic As in the LM tailing with amendments was

significantly higher than that in the HM due to the magnetite induced As adsorption capacity.

Because of the presence of elevated soluble inorganic As species, the concentrations and the

proportion of As (III) in LM tailings were higher than those of the HM tailings under the

prevailing redox conditions in the well watered pots. The findings indicated the important

role of Fe-oxides (albeit crystalline magnetite) in the speciation of inorganic As (III) and As

(V) in the pore water of the Cu-tailings under amendment and revegetation. The Fe oxide

particles preferentially adsorb As (V) species from the pore water compared to As (III) (Jenne

et al. 1979; Kocar et al. 2006).

Both the redox conditions and the formation of As-adsorbing Fe oxyhydroxides on the

surfaces of magnetite particles were mediated by microbial processes. The minimal As (V)

y = 18.044e0.0155x

R² = 0.8957

0

300

600

900

1200

0 50 100 150 200 250 300 350

Tota

l A

s co

nce

ntr

atio

n (

µg L

-1)

Fe concentration (mg L-1)

73

transformation into other forms may be expected at neutral and aerobic conditions and in the

presence of Fe oxyhydroxides (Al-Abed et al. 2007). At aerobic conditions (200-500mv),

Masscheleyn (1991) found that the dominant As soluble species was As (V). In contrast, Ryu

(2002) found that As (III) was the primary As soluble species at reducing conditions (-170

mv) in the Owens Dry Lake, California. Similarly, Mitchell (2006) conducted a column

experiment and found that As (III) desorption from ferrihydrite was more than twice than that

of As (V) at neutral pH conditions. In pore water, microorganisms may catalyse As (V)

reduction into As (III) if the As (V) is not re-absorbed on the Fe and Al oxides after

dissolution from primary minerals in the pore water (Oremland and Stolz 2003; Stolz and

Oremland 1999; Zobrist et al. 2000). In this study, the high magnetite in HM tailings would

have re-adsorbed soluble As from the pore water and transfer As (III) to As (V). The

presence of high levels of Fe oxyhydroxides in the tailings may provide early protection of

As (V) through adsorption, thus decreasing the opportunity of its transformation into As (III)

by relevant bacteria. Su (2008) found that As (III) decreased up to 70% in solution with pure

magnetite at pH 7, because Fe3+

at the surface of magnetite may act as an oxidant to transfer

As (III) to As (V) in solution. Inorganic As such as As (III) and As (V) are considerably more

toxic than the organic As species monomethylarsonic acid (MMA) and dimethylarsenic acid

(DMA) (Kaltreider et al. 2001). The exact processes of As speciation in the LM and HM

tailings amended with organic matter still need to be clarified, despite the consistent

observations of altered As speciation patterns.

4.3.3 Arsenic uptake by native grass and implication for remediation

It is not surprising to have observed that the plants accumulated more As when grown in the

LM tailings amended with organic matter, because of the elevated soluble As concentrations

in the pore water and the much reduced As adsorption capacity in the solid phase (see

Chapter 3). The plant responses in As uptake and accumulation have further confirmed the

importance of magnetite in the Cu-tailings, in terms of its roles in regulating As dissolution

and mobility in the pore water. The Red Flinders grasses accumulated more As in the

biomass, especially in roots grown in the LM tailings, compared to those in HM tailings due

to a higher concentration of bioavailable As (total and inorganic As) in pore water. The ratio

of As accumulation in shoots to roots was 60-86 and 14-19 in LM and HM tailings amended

with SR plus BC treatments. The high As accumulation in the roots may be caused by the

high Fe levels in roots which were up to 21 g kg-1

dwt in the two tailings, probably due to the

formation of Fe-plaque surrounding the root surface in the present study. This results in a

74

strong retention of As by the root surface and difficulty in estimating the actual amount of As

absorbed into the root cells (Figure 4-7) (Zhao et al. 2010). Therefore, the real As uptake by

roots may be lower than the results reported here.

The As bioaccumulation factor of Red Flinders grass was very low, 0.16 and 0.008 in the LM

and HM tailings, respectively. The low rate of As transfer from root to shoot was also

observed in native Spain Flora grown in the As-contaminated mined land (Moreno-Jiménez

et al. 2010). From the present findings, Red Flinders grass may be tentatively considered as

an “As excluder” under the biogeochemical conditions in the EHM tailings, although long-

term field trials are required to verify this speculation. Therefore, in future remediation

practices, the Red Flinders grass is suitable for phytostabilization in EHM tailings. According

to the National Research Council (2005), the maximum tolerable levels of As consumption

(MTL) for grazing animals (such as cattle, goats) ranges from 30 to 100 mg kg-1

dwt. If the

Red Flinders grasses grown in the LM (1.2-1.4 and 72-121 mg kg-1

dwt, in shoot and root,

respectively) HM tailings (0.7-1.0 and 10-15 mg kg-1

dwt in shoot and root, respectively), its

shoot As concentrations may not exceed the stipulated As levels for grazing animals.

Figure 4-7: Iron concentration in the root of Red Flinders grass harvested at the 4th

week after

commencing treatments in the organic matter amended LM and HM tailings

a a

b

a b ab

0

5

10

15

20

25

30


Fe

conce

ntr

atio

n (

g k

g-1

dry

wei

ght)

Treatment

HM tailings

LM tailings

75


compare the differences between treatments and different letters indicate their significant


4.4 Summary

The As bioavailability in the pore water was greatly elevated in the LM tailings compared to

the HM tailings under organic matter amendment and revegetation. This resulted from the

much reduced As adsorption capacity in the LM tailings, which rendered the increased

available As substrated in the process of bacteria-mediated speciation from inorganic to

organic forms in the solution phase. The organic matter amendments (mostly from sugarcane

residue) enhanced As at a much higher rate of As dissolution in the LM tailings than that of

the HM tailings. Without organic matter amendment, arsenic leaching into pore water was

highly limited in the tailings. The magnetite in tailings could act as an As sink in two ways:

(1) the magnetite itself has a capacity for As adsorption; (2) during the magnetite

transformation by reducing/oxidising bacteria, amorphous Fe minerals such as ferrihydrites

and siderite could be formed and precipitate on the surfaces of magnetite and arsenopyrite

particles, thus presenting a high As adsorption potential in the Cu tailings.

This differential As adsorption capacity in the LM tailings provides increased opportunities

of As transformation from inorganic into organic forms by bacteria present in the amended

tailings. On the basis of this limited study, the Red Flinders grass may be useful for

phytostabilizing the tailings surfaces with limited As accumulation in the shoots which may

be grazed by domestic and wild animals.

76

Chapter 5 General Discussion

The present findings have demonstrated that changes in ore processing technology such as

the recovery of magnetite after Cu-flotation could have significant environmental

consequences regarding the mobilisation of metalloids including As in the Cu-tailings. The

magnetite removal process and organic matter amendments change the tailings properties,

resulting in an altered distribution of As chemical forms in the solid and solution phases. This

can have flow-on effects on As concentrations in the pore water (and the resulting plant

uptake) when the tailings are subject to phytostablization. Therefore, understanding the

consequences facilitates informed decisions and practices when designing management

options for the LM tailings before and after revegetation, in particular in regards to seepage

water resulting from the tailings profile and eco-toxicity of pasture herbage.

Many biogeochemical processes may be involved in the continuum of mineral dissolution

and transformation. Many of these processes impact on As solubility and speciation in the

Cu-tailings subject to organic matter amendment and revegetation. Based on the present

findings and literature review, three possible mechanisms involved in the dissolution,

distribution and speciation of As in the tailings have been proposed (Figure 5-1) in regards to

its response to magnetite contents and organic matter amendments. These include:

(1) Mechanism (1) – enhanced As-adsorption capacity at the magnetite surface through

Fe dissolution and formation of amorphous Fe minerals which precipitate on the

particle surfaces.

(2) Mechanism (2) – increased labile carbon supply in the tailings may stimulate the

activity of reducing and oxidising bacteria and the dissolution of As/Fe-minerals to

release soluble As and Fe, which are then re-adsorbed by primary and secondary Fe-

minerals present or newly formed, leading to altered chemical forms of As in the

tailings.

(3) Mechanism (3) – the conversion of inorganic As into organic As may be increased by

the reduced As-adsorption capacity in the tailings, which may be further enhanced by

increasing available organic carbon supply when the tailings are amended with

organic matter for revegetation purposes.

77

5.1 Major differences in As fractionation between LM and HM Cu tailings.

Arsenic adsorption and distribution in tailings are closely related to the Fe-bearing minerals

and biogeochemical conditions which influence the dissolution and transformation of these

minerals (Pantsar-Kallio and Manninen 1997; Redman et al. 2002) (Chapter 3). The

magnetite removal process at EHM directly reduced the As adsorption capacity in the tailings

(Chapter 3). The high magnetite (HM) tailings had a stronger As (V) adsorption capacity than

the low magnetite (LM) tailings, even though the total As concentration in the LM tailings

was also lower than that in the HM (Chapter 3). The altered As-adsorption capacity by the

magnetite recovery process caused changes in As distribution patterns with much reduced

distribution in the specifically (exchangeable) and amorphous Fe phases in the LM magnetite

tailings (due to the lower contents of Fe-minerals) (Chapter 3).

Although the direct As adsorption capacity of magnetite is weaker than amorphous Fe-oxides

and oxyhydroxides (Giménez et al. 2007; Mamindy-Pajany et al. 2011), the combination of

magnetite particle surface area and the resultant amorphous Fe-minerals deposited on the

particle surface can greatly enhance As-adsorption capacity in the tailings. As a result, the

surface modification of magnetite particles through Fe-dissolution and precipitation of

amorphous Fe minerals may be a critical mechanism in the regulatory role of magnetite in As

adsorption and distribution in the Cu-tailings under revegetation (Figure 5-1 (1)). This may

be particularly enhanced when the tailings are amended with organic matter for revegetation

purposes under well-watered conditions. Organic matter amendment and revegetation greatly

increase the dissolution of Fe minerals as indicated by the elevated soluble Fe concentrations

in both the LM and HM tailings (Chapter 3).

The distribution of Fe forms has been altered in the tailings under direct revegetation

treatments, resulting in increased extractable amorphous Fe, compared to the unamended

treatment in both the LM and HM tailings (Chapter 3). It is interesting that the Fe

concentration in the pore water appeared much higher than in the HM tailings. It is possible

that the LM tailings under organic matter amendment and revegetaiton may present more

reducing conditions than the HM because of the finer particle size distribution in the LM after

further grinding and magnetite recovery. However, the measurement of redox potentials in

the freshly collected leachate samples did not reveal great differences between the LM and

HM tailings. To overcome these shortcomings in measurement, in situ monitoring of the

redox changes in the tailings may help to reveal the differences in redox conditions between

78

the LM and HM tailings under revegetation conditions. It is unclear whether differential

microbial processes were also involved in the increased Fe dissolution in the LM tailings; this

should be characterised in future studies. The relationships between the reducing conditions,

the activity of Fe-reducing/oxidising bacteria and the rate and quantity of ferrihydrides and

other secondary amorphous Fe minerals in the tailings of different magnetite contents under

OM amendments are also unclear.

The organic matter, for example sugarcane residue, may have increased the activity of Fe

oxidising and reducing bacteria in the tailings, resulting the dissolution of As/Fe-containing

primary minerals and releasing associated As into the pore water (Figure 5-1(2)). The iron

reducing bacteria Shewanella putrefaciens could use the Fe (III) on the surface of the

magnetite for metabolism and may form a layer (which may contain other iron compound

such as Fe (OH)2 and amorphous ferrihydrite) coating on magnetite surface. The layer can re-

adsorbed the As extracting from the amorphous iron phases (Roden and Zachara 1996). The

amorphous and crystalline Fe results (Chapter 3) in the tailings were consistent with the

distribution of As associated with the amorphous Fe minerals, which significantly increased

in the organic matter amended tailings, compared the unamended treatment. A highly and

positively correlated association was observed between concentrations of soluble As and Fe

in the pore water of the tailings amended with organic matter under revegetation.

5.2 Mechanisms of As dissolution and speciation in pore water and plants uptake

The major impact of magnetite recovery was found to be the much-increased concentrations

of total soluble As and organic As forms in the pore water of the LM tailings subject to

organic matter amendments (Chapter 4). This was in contrast to the greatly reduced total As

concentrations in the LM tailings (Chapter 3). The magnetite recovery process may have also

directly removed large proportions of As primary minerals arsenopyrite and/or soluble As

derived from the oxidation of arsenopyrite in the process of magnetite washing; arsenopyrite

is unstable and can be rapidly oxidised upon exposure to water and oxygen (Hernández and

Canadell 2008). The significantly elevated soluble As concentrations in the pore water of the

LM tailings may have been resulted from the combined effects of much reduced As

adsorption capacity and increased As-bearing mineral dissolution (Figure 5-1 (1) and (2)).

79

The high levels of soluble As in the pore water of the LM tailings provided a substrate for

bacteria-mediated As transformation or speciation into organic forms (Chapter 4). Under the

revegetation conditions in the glasshouse, the inorganic As, especially As (III), in the LM

tailings pore water was higher than that of HM tailings. This may be due to the prevailing

reducing conditions in the tailings under well-watered conditions. In addition, bacteria that

can use As (V) as the electron accepter for respiration in the amended tailings may have also

regulated the reduction of As (V) into As (III) (Oremland and Stolz 2003; Stolz and

Oremland 1999; Zobrist et al. 2000). The soluble As (III) and As (V) may be re-adsorbed

onto the surfaces of Fe and Al oxides such as in the case of the HM tailings, which may have

temporarily prevented access by As-respiring bacteria to carry out the conversion into organic

As (i.e. methylation process) (Figure 5-1 (2) & (3)). In addition, the reduction of As (V) into

As (III) may have decreased as the As (V) re-adsorbed to the surfaces of magnetite particles

forming secondary Fe minerals in the tailings.

Due to the elevated soluble As concentration present in the pore water of the LM tailings, the

native Red Flinders grass accumulated more As in the shoots compared to those of the HM

tailings. Within the plants, As was retained in roots (up to 146 mg kg-1

) more than shoots (2

mg kg-1

) probably because of Fe plaque at the root surfaces. Based on the bioaccumulation

factor, this grass can be regarded as the As-excluder species under the growth conditions,

rather than an As accumulating species. The plants may have also taken up some of the

organic As such as the MMA and DMA, but their uptake would be much lower than those of

the inorganic As species in most of plant species (Raab et al. 2007). Further field experiments

using a diverse range of native plant species should be tested to revegetate suitable native

pasture ecosystems and phytostabilize the tailings impoundment area with minimal risks of

As intake by grazing animals at the EHM.

5.3 Conclusions

In summary, the magnetite recovery resulted in significantly reduced As-adsorption capacity

in the Cu-tailings at EHM. Under the organic matter amendments, the levels of soluble As in

the pore water were greatly elevated in the LM tailings, due to the lack of adequate As-

adsorbing Fe-minerals including magnetite particles and newly formed amorphous Fe

oxyhydroxides which may have precipitated on the magnetite particles and other mineral

particles. This elevated soluble As in the LM tailings resulted in an increased uptake and

accumulation of As in Red Flinders grass, which was grown under well-watered conditions.

80

Research questions have been identified regarding detailed mechanisms in the discussion,

which will be investigated as part of a PhD project commencing in 2015.

(2) Dissolution-

adsorption

As solubility & speciation

As uptake and

accumulation in shoots

of pasture species

Organic matter

& plant species

Organic

As

Inorganic

As Food chain effects

in grazing animals

(1)Surface Fe dissolution

& modification

Adsorption

As (V) Fe-oxyhydroxides

& other secondary

minerals

Microbial

reduction/oxidation

Fe3O4

Low As (V) affinity High As (V) affinity

As-Fe

Minerals

Fe3O4

(2)Transformation

(3) As chemical forms Microbial

speciation

Figure 5-1: A conceptual diagram illustrating possible mechanisms of As and Fe mineral

dissolution, transformation, adsorption and speciation and plants uptake in Cu tailings under

organic matter amendment and revegetation. Three mechanisms have been proposed (1), (2) and

(3) in the diagram which has been interpreted in the discussion.

81

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Chapter 6 Supplementary Figures

Watering by

capillarity

\\

Amended tailings

7 cm

Capillary mat

Water level

Pore water sampler

Figure 6-1: The setup of twin-pot system and pore-water sampler

99

A hole of pore

water sampler

Roots in the amended tailings

Figure 6-2: Plant growth in the amended tailings by using the twin-pot system. Plants were

irrigated by bottom-fed water via capillary suction

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