THE USE OF THE INTEGRATED SOIL MICROCOSMS TO
ASSESS ACCUMULATION OF CAESIUM (Cs) AND LEAD (Pb)
FROM CONTAMINATED SOILS
BY EARTHWORMS (Eisenia andrei) AND
THE SUNFLOWER (Helianthus annuus)
HO VU VUONG
School of Applied Sciences
RMIT University, Melbourne, Australia.
March 2013
Master by Research Thesis
Declaration
I certify that except where due acknowledgement has been made, the work is that of
mine alone; the work has not been submitted previously, in whole or in part, to qualify
for any other academic award; the content of the thesis is the result of work which has
been carried out since the official commencement date of the approved research
program; and editorial work, paid or unpaid, carried out by a third party is
acknowledged.
HO VU VUONG
Date: 28 March 2013
ii
Acknowledgements
I would like to particular thank to my supervisor Professor Dayanthi Nugegoda and
co-supervisor Doctor José M. L. Rodrigues, for their continued and enthusiastic help
and support, encouragement, discussions, reading and correcting drafts and guidance
throughout my project.
I would like to thank RMIT University and the Ministry of Education and Training
Vietnam (MOET) in supporting me with the scholarship for my program. I would like
to forward my appreciation to Professor Andrew Smith, head of School of Applied
Sciences in giving me tuition fee for my extended semester. I also appreciate the
enthusiastic help from all staffs and officers of RMIT Vietnam and Australia during
my English courses and my Master program. I received a lot of in time help from my
first steps till the end.
I particular thank to Kathryn Thomas who helped me a lot with the paper works of the
scholarship for the enrolment, visa application and advices for my first days in
Australia. I also thank to Prof. Ngoc Tuan Nguyen, the coordinator of the program
with his enthusiastic help as a bridge connected with MOET.
I would like to thank to Professor Dinh Don Le who was my supervisor in Vietnam. I
thank to all staffs from Research Institute of Biotechnology and Environment, Nong
Lam University, Vietnam. I would also thank to all of my friends and colleges in
Vietnam, Australia and overseas (especially Thi Thu Hao Van and Ana Miranda) who
always stay by my side to give a hand and give me advices.
I would like to thank my grandparents, my parents, brothers, sisters and relatives for
their love, unconditional support and encouragement. Last but not least, thank you to
my beloved wife Thi Bich Thao Nguyen for her love, understanding and for being
always beside me.
iii
Table of contents
Declaration...................................................................................................................... ii
Acknowledgements ....................................................................................................... iii
Table of contents ............................................................................................................iv
List of figures............................................................................................................... viii
List of tables ....................................................................................................................x
Summary.........................................................................................................................xi
Common abbreviations ................................................................................................ xiii
List of communications .................................................................................................xv
Chapter 1. Introduction ....................................................................................................1
1.1Contamination of soil by trace metals ............................................................................... 1
1.1.1 Soil resources ............................................................................................................. 1
1.1.2 World-wide soil contamination ................................................................................. 1
1.1.3Sources of heavy metals in soil .................................................................................. 3
1.2 Caesium ............................................................................................................................ 7
1.2.1 Characteristics of Cs .................................................................................................. 7
1.2.2 Sources of Cs pollution ............................................................................................. 8
1.2.3 Caesium in soils ......................................................................................................... 9
1.2.4 The Chemical behavior of Cs in soils...................................................................... 11
1.2.5 Caesium and health effect in humans ...................................................................... 12
1.2.6 Caesium and environmental effects ......................................................................... 12
1.2.7 Remediation of Cs polluted soils ............................................................................. 13
1.3 Lead ................................................................................................................................ 13
1.3.1Characteristics of Pb ................................................................................................. 13
1.3.2 Sources of Pb ........................................................................................................... 14
1.3.3 Lead in soils ............................................................................................................. 15
1.3.4 The chemical behavior of Pb in soil ........................................................................ 16
1.3.5 Lead and health effects in humans .......................................................................... 17
1.3.6 Lead and environmental effects ............................................................................... 17
1.3.7 Remediation of Pb polluted soils ............................................................................. 18
1.4 Remediation of contaminated soils ................................................................................ 19
1.4.1 Remediation of heavy metals in soils ...................................................................... 20
1.4.2 Remediation of radionuclides in soils ..................................................................... 23
iv
1.4.3 Phytoremediation of heavy metals and radionuclides ............................................. 24
1.5 Sunflowers for phytoremediation of trace metals in soils .............................................. 28
1.6 Earthworms for remediation of heavy metals and radionuclides. .................................. 29
1.7 Use of toxicity tests for environmental risk assessment of contaminated soils.............. 33
1.8 Integrated soil microcosms ............................................................................................. 35
1.9 Aims of the project ......................................................................................................... 36
Chapter 2. Materials and methods .................................................................................37
2.1. Soil collection ................................................................................................................ 37
2.2 Soil physical and chemical characterization ................................................................... 37
2.3 Integrated Soil Microcosms ............................................................................................ 39
2.4 Experimental design ....................................................................................................... 41
2.4.1 Experiment 1: Caesium ........................................................................................... 43
2.4.2 Experiment 2: Cs and Pbas a mixture of contaminants. .......................................... 43
2.5 Soil sampling, digestion and chemical analysis ............................................................. 44
2.6 Leachate assessment ....................................................................................................... 44
2.7 Earthworm sampling, digestion and chemical analysis .................................................. 44
2.8 Plant sampling, wet digestion and chemical analysis ..................................................... 45
2.9 Data analysis ................................................................................................................... 46
Chapter 3. Results ..........................................................................................................48
3.1 Physical and chemical characteristics of soil ................................................................. 48
3.2 Caesium experiment ....................................................................................................... 48
3.2.1. Soil parameters after 2 and 4 weeks in experimental ISMs ................................... 48
3.2.2 Caesium in soils after 2 and 4 weeks ....................................................................... 51
3.2.3 Effects of Cs on leachates ........................................................................................ 53
3.2.3.1 Leachate pH ...................................................................................................... 53
3.2.3.2 Caesium concentration in leachates .................................................................. 54
3.2.4 Effects of Cs on earthworms ................................................................................... 55
3.2.4.1 Earthworm survival .......................................................................................... 55
3.2.4.2 Effects of Cs on earthworm distribution .......................................................... 56
3.2.4.3 Effects of Cs on earthworm biomass ................................................................ 58
3.2.4.4 Caesium accumulation in earthworms .............................................................. 60
3.2.4.5 Caesium transfer factor (TF) ............................................................................ 63
3.2.4.6 Total uptake of Cs by earthworms .................................................................... 65
3.2.5 Effects of Cs on plants ............................................................................................. 67
3.2.5.1 Effects of Cs on plant germination rate ............................................................ 67
3.2.5.2 Effects of Cs on plant growth (shoot length) .................................................... 68
3.2.5.3 Effects of Cs on plant biomass ......................................................................... 69
v
3.2.5.4 Caesium accumulation by plants ...................................................................... 70
3.2.5.5 Caesium transfer factor (TF) in plants .............................................................. 73
3.2.5.6 Comparison of shoot and roots -translocation factor........................................ 74
3.2.5.7 Caesium total uptake in plants .......................................................................... 75
3.3 Lead and Caesium+Lead as a mixture experiment ......................................................... 76
3.3.1 Soil parameters after the 4 week experiment........................................................... 76
3.3.2 Lead concentration in soils ...................................................................................... 77
3.3.3 Caesium concentration in soils ................................................................................ 78
3.3.4 Potassium concentration in soils.............................................................................. 79
3.3.5 Effects of Pb only and Cs+Pb on leachates ............................................................. 80
3.3.5.1 pH of leachates ................................................................................................. 80
3.3.5.2 Lead concentration in leachates ........................................................................ 81
3.3.5.3 Caesium and potassium concentration in leachates .......................................... 82
3.3.6 Effects of Pb only and Cs+Pb as a mixture on earthworms .................................... 83
3.3.6.1 Effects of Pb only and Cs+Pb on earthworm survival ..................................... 83
3.3.6.2 Effects of Pb only and Cs+Pb on earthworm distribution ................................ 84
3.3.6.3 Effects of Pb only and Cs+Pb on earthworm fresh biomass ............................ 85
3.3.6.4 Lead accumulation in earthworms .................................................................... 86
3.3.6.5 Lead transfer factor in earthworms .................................................................. 87
3.3.6.6 Lead total uptakein earthworms ....................................................................... 88
3.3.6.7 Caesium in earthworms .................................................................................... 89
3.3.6.8 Caesium transfer factor in earthworms ............................................................. 90
3.3.6.9 Caesium total uptake in earthworms ................................................................ 91
3.3.6.10 Potassium in earthworms ................................................................................ 91
3.3.7 Effects of Pb only and Cs+Pb as a mixture on plants.............................................. 92
3.3.7.1 Plant germination .............................................................................................. 92
3.3.7.2 Plant growth (shoot length) .............................................................................. 93
3.3.7.3 Above ground dried biomass of plants ............................................................. 94
3.3.7.4 Pb accumulation in plants ................................................................................. 94
3.3.7.5 Lead transfer factor in sunflowers .................................................................... 96
3.3.7.6 Lead translocation factor in sunflowers............................................................ 97
3.3.7.7 Total uptake of Pb in sunflowers ...................................................................... 97
3.3.7.8 Caesium accumulation by plants ...................................................................... 98
3.3.7.9 Caesium transfer factor of sunflowers .............................................................. 99
3.3.7.10 Caesium translocation factor .......................................................................... 99
3.3.7.11 Total uptake of Cs in sunflowers .................................................................. 100
3.3.7.12 Potassium in plants ....................................................................................... 101
3.4 Estimation of the time required for effective phytoremediation .................................. 101
vi
Chapter 4. Discussion ..................................................................................................103
4.1 Integrated soil microcosms ........................................................................................... 103
4.2 Change of soil properties and metals in soil during the experiments ........................... 104
4.2.1
Caesium in soils:............................................................................................... 105
4.2.2
Lead in soils:..................................................................................................... 107
4.3 Leachates ...................................................................................................................... 108
4.4 Earthworms ................................................................................................................... 108
4.4.1 Mortality and mobility of earthworms .................................................................. 108
4.4.2 Caesium and Pb accumulation by earthworms ...................................................... 110
4.5 Plants ............................................................................................................................ 111
4.5.1 Plant germination, growth and biomass. ............................................................... 111
4.5.2 Caesium and Pb accumulation by plants: .............................................................. 113
4.6 Remediation efficiency ................................................................................................. 116
5. Conclusions .............................................................................................................118
6.
Future work ..........................................................................................................120
References ...................................................................................................................121
vii
List of figures
Figure
Page
Figure 1.1 Soil classifications ....................................................................................... 1
Figure 1.2 Dependence of the integrated activity concentration of radionuclides
during March and April 2011 on the distance from Fukushima ................................. 10
Figure 1.3 Comparison between Third Airbone Monitoring Results and Map of
Cs-134 Concentration ................................................................................ 11
Figure 1.4 Sunflowers (Helianthus annuus)................................................................ 29
Figure 1.5 Eisenia andrei ............................................................................................ 33
Figure 1.6 Main components of an Integrated Soil Microcosm (ISM) ....................... 35
Figure 2.1 Integrated soil microcosms ........................................................................ 40
Figure 2.2 Earthworm Eisenia andrei ......................................................................... 40
Figure 2.3 Sunflowers (Helianthus annuus)................................................................ 41
Figure 3.1 Cs concentrations in soils after 2 and 4 weeks .......................................... 52
Figure 3.2 Caesium concentrations in leachates after 2 and 4 weeks ......................... 55
Figure 3.3 Earthworm survival after 2 and 4 weeks .................................................... 56
Figure 3.4 Effects of Cs on earthworm distribution after 2 and 4 weeks .................... 57
Figure 3.5 Effects of Cs on earthworm biomass ......................................................... 59
Figure 3.6 Concentration of Cs in earthworms and soil after 2 and 4 weeks.............. 62
Figure 3.7 Caesium transfer factor in earthworms after 2 and 4 weeks ...................... 63
Figure 3.8 Total uptake of Cs by earthworms after 2 and 4 weeks ............................. 65
Figure 3.9 Effects of Cs on plant germination rate ..................................................... 67
Figure 3.10 Effects of Cs on shoot length after 2 and 4 weeks ................................... 68
Figure 3.11 Effects of Cs on above ground dried biomass after 2 and 4 weeks ......... 69
Figure 3.12 Concentration of Cs in plants after 2 and 4 weeks................................... 72
Figure 3.13.Caesium transfer factor in plants after 2 and 4 weeks ............................. 73
Figure 3.14 Caesium translocation factor in plants after 2 and 4 weeks ..................... 74
Figure 3.15 Caesium total uptake in plants after 2 and 4 weeks ................................. 75
Figure 3.16 Lead concentration in soils after 4 weeks ................................................ 78
Figure 3.17 Caesium in soils after 4 weeks ................................................................. 79
Figure 3.18 Potassium concentrations in soils after 4 weeks ...................................... 79
viii
Figure 3.19 Lead concentrations in leachates after 4 weeks ....................................... 81
Figure 3.20 Caesium concentration in leachates after 4 weeks ................................... 82
Figure 3.21 Potassium concentrations in leachates after 4 weeks ............................... 83
Figure 3.22 Effects of Pb and Cs+Pb on earthworm survival after 4 weeks............... 83
Figure 3.23 Effects of Pb and Cs+Pb on earthworm distribution after 4 weeks ......... 85
Figure 3.24 Effects of Pb only and Cs+Pb on earthworm biomass after 4 weeks ...... 85
Figure 3.25 Lead concentration in soil and earthworms after 4 weeks ....................... 86
Figure 3.26 Lead transfer factor in earthworms after 4 weeks .................................... 88
Figure 3.27 Lead total uptake in earthworms after 4 weeks ....................................... 88
Figure 3.28 Caesium concentration in earthworms after 4 weeks .............................. 89
Figure 3.29 Caesium transfer factor of earthworms after 4 weeks ............................. 90
Figure 3.30 Caesium total uptake in earthworms after 4 weeks ................................. 91
Figure 3.31 Potassium concentration in earthworms after 4 weeks ............................ 92
Figure 3.32 Effects of Pb only and Cs+Pb on plant germination rate ......................... 93
Figure 3.33 Effects of Pb only and Cs+Pb on plant growth (shoot length)
after 4 weeks ................................................................................................................ 93
Figure 3.34 Effects of Pb only and Cs+Pb on plant biomass (above ground dried
biomass) after 4 weeks ................................................................................................ 94
Figure 3.35 Lead concentrations in plants after 4 weeks ............................................ 95
Figure 3.36 Lead transfer factor in plants after 4 weeks ............................................. 96
Figure 3.37 Lead translocation factor in plants after 4 weeks..................................... 97
Figure 3.38 Lead total uptake in sunflowers after 4 weeks ......................................... 98
Figure 3.39 Caesium concentration in plants after 4 weeks ........................................ 98
Figure 3.40 Caesium transfer factor in plants after 4 weeks ....................................... 99
Figure 3.41 Caesium translocation factor of plants after 4 weeks ............................ 100
Figure 3.42 Caesium total uptake in plants after 4 weeks ......................................... 100
Figure 3.43 Potassium concentration in plants after 4 weeks ................................... 101
ix
List of tables
Table
Page
Table 1.1 The cleanup market sites in the United States............................................... 3
Table 1.2 Organic amendments for heavy metal immobilization ............................... 21
Table 1.3 Inorganic amendments for heavy metal immobilization ............................. 21
Table 1.4 Substances amenable to the phytoremediation process............................... 24
Table 2.1 Parameters measured at each sampling time ............................................... 42
Table 2.2 Experimental design used for Cs ................................................................. 43
Table 2.3 Cs and Pb accumulation experimental design ............................................. 44
Table 3.1 Properties of clean soil collected near the Ecotoxicology laboratory, RMIT
Bundoora west campus. ............................................................................................... 48
Table 3.2 Soil pH after 2 and 4 weeks in experimental ISMs ..................................... 49
Table 3.3 Soil conductivity (μS/cm) after 2 and 4 weeks in experimental ISMs........ 49
Table 3.4 Soil moisture (%) after 2 and 4 weeks in experimental ISMs ..................... 50
Table 3.5 Soil Organic Matter Content (%) after 2 and 4 weeks in experimental ISMs
..................................................................................................................................... 50
Table 3.6 pH of leachates after 2 and 4 weeks ............................................................ 53
Table 3.7 Soil pH after 4 weeks .................................................................................. 76
Table 3.8 Soil conductivity (μS/cm) after 4 weeks ..................................................... 76
Table 3.9 Soil moisture (%) after 4 weeks .................................................................. 77
Table 3.10 Soil organic matter content (%) after 4 weeks .......................................... 77
Table 3.11 Leachate pH after the four week experiment. ........................................... 80
Table 3.12 Estimation of time required for the phytoremediation of Cs and Pb using
sunflowers ....................................................................................................................102
x
Summary
Heavy metal and radionuclide contamination in soils from industrial and nuclear
activities is a serious problem and increasing. Most metals and metalloids (or
elements) present in contaminated soils in order of abundance are lead (Pb), chromium
(Cr), arsenic (As), zinc (Zn), cadmium (Cd), copper (Cu) and mercury (Hg); (USEPA,
1996). These elements are accumulated and magnified through the food chain and
could enter surface and groundwater. Of these, Pb is the most significant and
consistent contaminant in hazardous waste areas and found at more than 50% of the
contaminated sites of the NPL (National Priorities List) of the USA; while caesium is
commonly released in radioactive wastes (i.e.
137
Cs) and has contaminated large land
areas (Cambray et al., 1987, Helfand et al., 2002). Bioremediation is considered an
effective way to reduce metal contamination due to its low cost and environmental
sustainability. Phytoremediation is a bioremediation application which can be used to
decontaminate not only heavy metals (e.g. Ag, Cd, Co, Cr, Cu, Hg, Mn, Mo, Ni, Pb,
and Zn) but also radionuclides (e.g. 3H, 90Sr, 137Cs, 239Pu, 234U, 238U) in soils (Andrrade
and Mahler, 2002, Negri and Hinchman, 2000).
In this study the earthworm Eisenia andrei and sunflower dwarf sensation (Helianthus
annuus) were used as tools to accumulate Pb and stable Cs. The aim of this study was
to evaluate bioremediating Cs and Pb in contaminated soils using the selected biota.
The efficacy of this bioremediation was evaluated using integrated soil microcosms
(ISM)–a multispecies mini ecosystem with soil contaminated with Cs alone at 25 mg/kg
and 250 mg/kg; Pb alone (1500 mg/kg) and Pb (1500 mg/kg) together with Cs (250
mg/kg) in the laboratory. Results revealed that the sunflower was able to
bioaccumulate Cs and Pb from contaminated soil and that Cs and Pb were
accumulated more in roots than in the shoots of the plants. The Cs bioaccumulated in
plants increased when the concentration of Cs in soils was higher. The presence of
earthworms did not increase Cs and Pb accumulation or the transfer of the trace metals
from roots to shoots in the plants. The presence of Cs+Pb as a mixture reduced the
accumulation, transfer factor from soil to plants and the translocation of Cs and Pb.
Earthworms had a high tolerance for Cs and Pb and accumulated a very high level of
xi
Cs (250 mg/kg) and Pb (1500 mg/kg), and the accumulated concentrations increased
when the concentration of trace metals in the soil was higher. It was also evident that
the Pb in earthworms in the top layer of the soil core was higher than in those in the
lower core. The presence of Pb with Cs as a mixture reduced the accumulation of both
Pb and Cs in earthworms. Lead was only marginally removed from the soil in the
treatments with only 0.9 to 5.5% decontamination after treatment with plants and
earthworms while Cs was removed from soil to a greater extent, estimated here to be
27 to 49%, proportional to concentration of trace metal in the contaminated soils. Lead
and caesium were also detected in leachates however the concentration was very low
and below the limit of Australian drinking water guidelines for Pb and well below the
ANZECC water quality guideline for Pb .The study also demonstrated that Pb in soil
was less mobile and less bioavailable than Cs.
xii
Common abbreviations
AAS
Atomic Absorption Spectrometry
AFCF
Amonium-Ferric-Hexacyano-Ferrate (II)
ANOVA
Analysis Of Variance
ASTM
American Society for Testing and Materials
BF
Bioaccumulation factor
CSTEE
EU’s scientific committee on environmental toxicology
DNA
Deoxyribonucleic acid
DMT
N,N-Dimethyltryptamine
DOC
Dissolved Organic Carbon
DOE
Department Of Energy
DTPA
Diethylene Triaminepenta Acetic Acid
EDDS
Ethylene Diamine Disuccinate
EDGA
Glycoether Diamine Tetraacetic
EDTA
Ethylene Diamine Tetraacetic
FAO
Food and Agricultural Organization
GWRTAC
Ground Water Remediation Technology Analysis Center of USEPA
HA
Humic acids
HDPE
High-Density Polyethylene
HDR
Higher Degree Research
ICP-MS
Inductively Coupled Plasma-Mass Spectrometry
INEEL
Idaho National engineering and environmental laborator
IPCS
International Programme on Chemical Safety of WHO
ISM
Integrated Soil Microcosm
ISO
International Organization for Standardization
JAEA
Japan Atomic Energy Agency
LMWOAs
Low Molecular Weight Organic Acids
LOI
Loss On Ignition
MEXT
Ministry of education, culture, sports, science and Technology,Japan
MTBE
Methyl Tertiary Butyl Ether
NOES
National Occupational Exposure Survey
NTA
Nitrilotriacetic Acid
xiii
OECD
Organization for Economical Cooperation and Development
OMC
Organic Matter Content
PAH’s
Polycyclic Aromatic Hydrocarbons
PCB
Polychlorinated Biphenyls
PCE
Perchloroethylene
PNEC
The predicted no-effect concentration
PVC
Polyvinyl Chloride
RDX
(C3H6N6O6) Hexahydro-1,3,5-Trinitro-1,3,5-Triazine
SETAC
Society of Environmental Toxicology And Chemistry
TCE
Tetrachloroethylene
TF
Transfer Factor
TNT
2,4,6-Trinitrotoluene
UK
The United Kingdom
US.ATSDR
Agency for Toxic Substances and Disease Registry of United State
US.CDC
Centers for Disease Control and Prevention of United State
USA
The United State of America
USEPA
United Stated Environmental Protection Agency
VEB
Vertical Engineered Barries
WHO/UNECE
World Health Organization/The United Nations Economic Commission
for Europe
xiv
List of communications
Poster and oral presentations
An oral presentation was given at the 2nd International conference on environmental
pollution, restoration, and management. (Hanoi University of Science, Loyola
University Chicago and SETAC Asia Pacific Geographic Unit Joint Conference), in
Ha Noi, Viet Nam, March 4-8, 2013, Topic: The use of the integrated soil microcosms
to assess accumulation of lead (Pb) from contaminated soil by earthworms (Eisenia
andrei) and the sunflower (Helianthus annuus- Dwarf sensation).
Poster presentation at RMIT University HDR (Higher degree of research)-2012
conference, in Melbourne, Australia, in October 2012, Topic: The use of the integrated
soil microcosms to assess accumulation of lead (Pb) from contaminated soil by
earthworms (Eisenia andrei) and the sunflower (Helianthus annuus- Dwarf sensation).
Poster presentation at School of Applied Sciences research day, RMIT University,
Melbourne, Australia, in June2011, Topic: The use of the integrated soil microcosms
to assess accumulation of heavy metals from contaminated sediments and biosolids by
earthworms and plants.
Awards
The People’s choice best poster prize in the Sustainability stream sponsored by the
College of Science, Engineering and health, RMIT University HDR-2012 conference,
October 2012.
The College of Science, Engineering and Health’s choice best poster presentation in
the Sustainability stream, sponsored by Fuji Xerox, RMIT University HDR-2012
conference, October 2012.
xv
Chapter 1. Introduction
1.1 Contamination of soil by trace metals
1.1.1 Soil resources
Soils have been subjected to anthropogenic influence before most other natural
resources. There are approximately 17,000 kinds of soil activities identified in the
United States and uncertified numbers worldwide (Lynden, 2004). Land use has been
categorized by the Food and Agriculture Organization into four types: farming land,
grass land (rangelands), forest and general use including unused land in urban areas
and waste and barren land
(Figure 1.1) (FAO, 1984).
In an ecosystem, soil
Forests
4108
31.2%
plays four fundamental
roles which are (1) a
physical
supporting
structure
plants
to
conduct photosynthesize,
General use
4373
33.2%
Rangeland
3180
24.2%
Cropland
1500
11.4%
(2) cooperate with the air
to
create
a
suitable
Figure 1.1 Soil classifications
(FAO,(3)
1984)
temperature for living organisms in the ecosystem,
absorb, store and release water
to support plants in dry situations and (4) provide essential elements for plants,
animals and humans (Buol, 1995). Soil is an essential and non-renewable resource and
has not received much attention compared with air and water (Lynden, 2004).
1.1.2 World-wide soil contamination
The contamination of soils by heavy metals and radionuclides spreads across large
areas. In soil, contaminants exist in solution, as inorganic or inorganic complexes, and
as pure or mixed precipitated solids (Smith et al., 1995). Among heavy metal
contaminants, lead (Pb), chromium (Cr), zinc (Zn), arsenic (As), and cadmium (Cd)
are common in decreasing order in soil and groundwater (National research council,
1
1994, Knox et al., 1999). Heavy metals cannot be biodegraded with time while organic
and radionuclide contaminants are degradable and subject to half life reduction
respectively. Heavy metal and radionuclides contamination in soils is raising concern
on the loss of ecosystem services, agricultural productivity, food chain quality
reduction, groundwater pollution, economic loss, and human and animal diseases.
(Dushenkov, 2003, Endo et al., 2012, Adriano et al., 1997)
In the U.S, according to a report in 1997, there was an estimation of nearly half a
million contaminated sites and more than 217,000 of them need to be cleaned up. The
cleanup market showed in table 1.1 includes the Environmental Protection Agency’s
(EPA) Superfund sites and the Resource Conservation and Recovery Act (RCRA),
Department of Defense (DOD), Department of Energy (DOE), State sites, and Private
Party sites.
In Europe, the United Kingdom contains 325,000 potentially contaminated sites (2%
of land area) equal to 300,000 ha. 33,500 contaminated sites were identified to date.
21,000 sites have been treated. In Germany, 360,000 potential contaminated sites were
determined by the German Federal Ministry for the environment, Nature Conservation
and Nuclear Safety, 2002. In Denmark, 40,000 potential contaminated sites were
found by the Danish Ministry of the Environment, 2009. In the Netherlands, 60,000
potential contaminated sites were recorded by the Dutch Ministry of Housing, Spatial
Planning and the Environment (now part of the Ministry of Infrastructure and
Environment), 2009. (Mark, 2012). In Australia, nearly 80,000 contaminated sites
were identified up to 2008 (Naidu et al., 2008).
2
Table 1.1 The cleanup market sites in the United States
(Raskin and Ensley, 2000).
Cleanup market share in the United States
Organic and heavy
Heavy metal
metal contaminated
contaminated
Throughout the U.S
217, 000 sites
Environmental Protection Agency’s (EPA) and
The Resource Conservation and Recovery Act
64%
15%
(RCRA) sites
Department of Defense (DOD) sites
11%
(7313=26,000 acres)
Department of Energy (DOE) sites (4000 (23
53%
7%
38%
7%
listed as Superfund sites)
State sites (19,000)
Private Party sites
24,000 sites
1.1.3 Sources of heavy metals in soil
Soils receive heavy metals naturally from the pedogenetic process with low toxicity
(Pierzynski et al., 2000, Kabata-Pendias and Pendias, 2001). However, human
activities have intervened to accelerate and increase the concentration of heavy metals
in rural and urban areas to a concentration that can be harm to the human health,
animals, plants and the ecosystem. Heavy metal toxicity in the environment is due to
their production rate increasing by human activities, the transfer from mines to other
areas that can pose a greater potential of harmful exposure, their content in wastes
being higher than those in the deposit areas, and metals becoming more bioavailable in
the receiving environment (D'Amore et al., 2005). A simple equation for the total
amount of heavy metal in soil can be figured out as below: (Alloway, 1995, Lombi and
Gerzabek, 1998)
Mtotal=(Mp+Ma+Mf+Mag+Mow+Mip)-(Mcr+Ml)
Where “M” is the heavy metal, “p” is the parent material such as rock, “a” is
the atmospheric deposition, “f” is the fertilizer sources, “ag” are the agrochemical
3
sources, “ow” are the organic waste sources, “ip” are other inorganic pollutants, “cr” is
crop removal, and “l” is the loss by leaching, volatilization, and so on.
Soil acts as a sink for heavy metals through the human activities mentioned above.
Metals cannot be degraded by microbial or chemical pathways (Kirpichtchikova et al.,
2006) and exist for a long time (Adriano, 2003). They also inhibit the biodegradation
of organic contaminants (Maslin and Maier, 2000). Heavy metals are considered
inorganic hazards causing diseases, and among these lead (Pb), chromium (Cr), arsenic
(As), zinc (Zn), cadmium (Cd), copper (Cu), mercury (Hg), and nickel (Ni) are the
most notable. They may cause harm to humans and the ecosystem following direct
consummation or contact via food chains (soil-plant-human or soil-plant-animalhuman), polluted drinking ground water, food quality and security reduction (Wuana
and Okieimen, 2011).
Soil can be polluted by the entering of heavy metals and metalloids from industrial
areas, mining, high metal waste disposal, Pb containing fuel and paints, fertilizers,
animal wastes, sewage sludge, pesticides, herbicides, wastewater application in
agriculture, burned residues of coal, petrochemicals spills, and natural deposition
(Wuana and Okieimen, 2011).
Agriculture was the first impact of humans on the soil (Scragg, 2006). In order to
support plant growth, both macronutrients (N, P, K, S, Ca, and Mg) and trace-elements
(Co, Cu, Fe, Mn, Mo, Ni, and Zn) (Lasat, 2000) were added to the soils or sprayed
onto it. Additional copper can be introduced into low copper soils to enhance the
growth of cereal crops, and Mn can also often be added into cereal and root crops
soils. A huge amount of fertilizers have been being introduced to soils in intensive
agriculture to guarantee an adequate N, P, and K. These fertilizers often contain toxic
heavy metals such as cadmium (Cd) and Pb as impurities which may be considerably
increased in agricultural soil by long term application (Jones and Jarvis, 1981), (Raven
et al., 1998).
4
Some well-known pesticides used in agriculture and horticulture contain high
concentration of heavy metals. For example, approximately 10% of the substances that
have been approved as insecticides and fungicides in UK were chemicals that contain
Cu, Hg, Mn, Pb, or Zn. Bordeaux mixture (copper sulphate) and copper oxychloride
are copper-containing fungicidal sprays (Jones and Jarvis, 1981). Lead arsenate was
another parasitic insecticide used in fruit orchards for long time. In New Zealand and
Australia, arsenic containing substances were also used widely to deal with cattle ticks
and pests in banana. Combinations of Cu, Cr, and As (CCA) compounds have been
used to preserve timbers. Now, there are many contaminated sites where the
concentration of these elements exceeds the environmental quality guidelines. The
contamination is raising the potential risk to human health from trace metals if these
sites reused for agriculture or nonagricultural purposes (McLaughlin et al., 2000).
The wide application of biosolids (e.g., livestock municipal, composts, and sewage
sludge) on landfill can cause the great increases of heavy metals such as arsenic (As),
cadmium (Cd), chrome (Cr), copper (Cu), lead (Pb), mercury (Hg), nickel (Ni),
selenium (Se), molybdenum (Mo), zinc (Zn), thallium (Tl), antimony (Sb) etc. in soil
(Basta et al., 2005). Particular manures of poultry, cattle and pig are applied on crops
and pastures as solid or slurries and are considered valuable fertilizers. However, in
the food of pigs and poultry, copper and zinc are introduced to increase the growth rate
and arsenic included in poultry health chemicals. These can cause the heavy metal
contamination in soil as well as gradually increase their concentration if manures from
these sources are used frequently in a certain areas (Sumner, 2000, Chaney and Oliver,
1996).
Sewage sludge from wastewater treatments has been commonly used for fertilizer in
agriculture. In the United State, an estimation of approximate 2.8 million dry tones are
reused or disposed. In Europe, more than 30% sewage sludge was utilized in
agriculture (Silveira et al., 2003). In Australia, over 175,000 tones of biosolids are
used for crops. These can lead to the increase of heavy metals in soil (McLaughlin et
al., 2000). In addition, composting biosolids which originated from sawdust, straw and
garden wastes can be of potential harm to the soil because they are frequently applied
5
and contain heavy metals (Canet et al., 1998). Biosolid application under certain
conditions can lead to leaching of heavy metals to groundwater through the soil. For
example, studies have demonstrated the increasing level of Cd, Ni, and Zn in leachates
from biosolids applied to soils in New Zealand (Keller et al., 2002, McLaren et al.,
2004).
Farmers have been using municipal, industrial wastewater and other effluents. It is
estimated that approximately 20 million hectares worldwide are irrigated with
wastewater. In some Asian and African cities, 50% of vegetable cultivation areas are
treated with wastewater. The farmers are unaware of the consequence of using these
sources to the environment but interested in yielding their products instead. Permanent
irrigation will lead to soil accumulation although the concentrations of heavy metals in
these sources are low (Bjuhr, 2007).
Mining and milling activities with resulting industrial wastes have been serious
problems in many countries around the world. The wastes during the mining process
are disposed directly into the environment. This result in the elevation of heavy metals
then poses risks to humans and the ecosystem. The soil productivity may not be
restored even using some long-term and expensive technologies (DeVolder et al.,
2003, Basta and Gradwohl, 1998). Heavy metals (Cr, Pb, and Zn) are also released
from industries such as textile, tanning, petroleum chemicals or petroleum-based
products, pesticides and pharmaceutical (Sumner, 2000).
Airborne sources of heavy metals are of different types including stack or duct
emissions through air, gas or vapor streams and fugitive emissions from storage areas
or waste piles. Generally, metals from airborne sources are disposed of in the gas
stream. Some heavy metals such as As, Cd, and Pb can be evaporated at high
temperatures. They will become oxides and condense as particles in an oxidizing
atmosphere (Smith et al., 1995). Stack emission can be spread in a large area by the
wind. They can only be removed from the air flow by dry and/or wet precipitation. In
contrast, fugitive emissions are present in a smaller area and close to the ground. The
forms and concentrations of metals in these sources will depend on the condition of the
6
site emitted. All particles from fire and factory chimneys will be deposited on land and
oceans; fossil fuels contain some heavy metals and therefore this was the major source
of contamination on a large scale following by the industrial revolution. For instance,
high concentrations of Cd, Pb, and Zn have been found in plants and soils next to
smelting factories. Using tetraethyl Pb petrol released Pb which contaminated soils in
urban areas and soils at road sites; tyres and lubricant oil can also produce Zn and Cd
(USEPA, 1996).
The most common heavy metals present in contaminated soils in order of abundance
are Pb, Cr, As, Zn, Cd, Cu and Hg (USEPA, 1996). They are noteworthy because they
are accumulated and magnified through the food chain. They also infiltrate to surface
and groundwater (Levy et al., 1992). Of these, Pb is the most significant contaminant.
Lead contamination was found at more than 50% of the NPL (National Priorities List)
of the USA. Lead is one of the most consistent contaminants in hazardous waste areas.
Stable Cs has not commonly been considered as a metal of concern in contaminated
sites. As a result there are few previous studies on the bioremediation of Cs. However
in view of the wide contamination by137Cs and 134Cs following the nuclear accidents in
Chernobyl (1986) in the Ukraine and Fukushima (2011) in Japan; it was considered of
interest to evaluate whether Cs could be bioremediated in contaminated soils in the
current research project. A mixture of Pb with Cs was also of interest since Pb is often
the end product following radioactive decay and maybe found in sites contaminated by
nuclear waste along with radiocaesium (Scotti and Carini, 2000). Hence the second
experiment in the current study evaluated bioremediation of Cs and Pb as a
contaminant mixture in soils.
1.2 Caesium
1.2.1 Characteristics of Cs
Caesium is a soft, ductile, and silvery white alkaline metal. It has one oxidation
isoform (+1). Caesium has a low melting point, it becomes liquid at slightly above
room temperature. Low boiling point and ionization potential, high vapor pressure, and
highest density are properties of Cs when compared with other stable alkaline metals.
7
Caesium is far more active than alkali metal members. Caesium produces a reddish
violet flame when exposed to air and forms caesium oxides. Similar to other alkali
metals, Cs strongly reacts with water and creates caesium hydroxide and hydrogen gas.
Its salts and compounds are soluble in water except caesium alkyl and aryl compounds
(ATSDR, 2004).
Isotopes of Cs range from
(114Cs).
135
114
Cs to
145
Cs. The shortest half life is about 0.57 seconds
Cs has longest half life with 3x106 years (Helmers, 1996).
137
Cs and
134
Cs
are widely used since they have high fission yield (137Cs produces 6 atoms per 100
fission events (WHO, 1983a) and relatively long half lives (about 30.2 and 2.1 years of
137
Cs and
134
Cs respectively). Stable Cs and radiocaesium have the same chemical
properties (ATSDR, 2004).
1.2.2 Sources of Cs pollution
133
Cs is a natural and stable isotope of Cs and Cs minerals. It is present at low
concentration in the earth’s crust (about 1 mg/kg in granites and 4 mg/kg in sediment
rocks) (Burt, 1993). Mineral pollucite containing 5-32% Cs2O is the major source of
commercial Cs. Two-third of the world’s supply comes from Canada. Dust and erosion
are the main sources of Cs in the environment. The other sources of Cs are mining of
pollucite ores and electronic and energy industries. Caesium is also measured in fly
ash from coal burning and waste incinerators (Fernandez et al., 1992, Mumma et al.,
1990).
There is more concern about radiocaesium (137Cs and
form. The average concentration of
et al., 1993).
134
Cs and
137
137
134
Cs) rather than the stable
Cs in the U.S soil was 620 mCi/km2 (Holmgren
Cs have occurred since 1945 by nuclear weapon testing,
nuclear plant wastes, and nuclear facility accidents (Avery, 1996).
In the 1950s and 1960s, the atmospheric wet and/ or dry depositions up to 1.5x1018Bq
of radiocaesium were determined to have originated from nuclear weapons testing.
Atmospheric fallout was the major sources of food chain contamination through
vegetation deposition (Holmgren et al., 1993, Meriwether et al., 1988). According to
the guideline for remediation strategies by the end of 1986, land with deposition
8
density higher than 3700 kBq/m2 was banned from agricultural production (Fesenko
and Howard, 2012).
In addition, radiocaesium fall out from nuclear plant accidents such as Chernobyl,
Ukraine (1986) (Mench et al., 2000) and Fukushima Dai-ichi, Japan (2011) (MEXT,
2011a) have resulted in radiocaesium pollution of large areas. The Chernobyl nuclear
accident polluted 6% of the European territory with
137
Cs at level above 20 kBq/m2.
Levels above 40 kBq/m2 and 1480 kBq/m2 were found in 2% and 0.03% in Europe
(Bruno et al., 2000).
Furthermore, radiocaesium exists in buried radioactive waste materials (Mench et al.,
2000, Iskandar, 1992, Meriwether et al., 1988). Caesium in wastes is considered as
intermediate in solubility.
1.2.3 Caesium in soils
Stable Cs (133Cs) occurs in soils through mining, milling, coal and burning of wastes.
In the bottom ash of municipal solid waste incinerators, a concentration of 0.44-2.01
mg/kg and 3-23 mg/kg of
133
Cs was found in the United States and in Spain
respectively (Fernandez et al., 1992). Stable Cs was detected both in soil and sediment
at one of the eight National Priorities Lists (NPL) hazardous waste sites in the U.S
(HazDat., 2003).
Radiocaesium has been introduced into soil by nuclear weapons testing (under and
upper ground) and nuclear plant accidents (Chernobyl, Fukushima Dai-ichi).
According to the Agency for Toxic Substances and Disease Registry (1993), about
1,400 underground testing of nuclear weapons have been conducted all over the world.
137
Cs and
134
Cs small release are also originated from nuclear plant general operation
due to the storage and use of radioceium (Radiation, 1996).A range of
137
Cs levels
from 1.6x10-8 to 3.4x10-7Ci/m2 was determined at a transuranic wastes storage site in
the US (the Idaho National Engineering and Environmental Laboratory (INEEL))
(DOE, 1998a). The average concentration of
137
Cs and
134
Cs in the top 0-8 cm soil
layer near the Chernobyl nuclear power plant, several years after the Chernobyl
9
accident was 8.6x10-5 Ci/m2 and 1.9x10-5 Ci/m2 respectively (Mikhaylovskaya et al.,
1993).
Nuclear accidents in Chernobyl, Ukraine (26 April 1986) and Fukushima, Japan led to
serious environmental concerns. The ecological damage was observed in a 30 km
radius from the reactor. After the Chernobyl accident more than 260,000 km2 were
exposed to higher than 1 Ci.Km-1 of 137Cs, received another 4.3 mSv of radiation level
equally to a 0.1% the risk of death cancer in Ukraine (Dushenkov, 2003, Endo et al.,
2012, Adriano et al., 1997). In addition, radiocaesium was identified in milk in the UK
in 1964 and 1986 compatible with the peak of nuclear weapons testing and the
Chernobyl accident, respectively (Department of the Environment, 1994).
The
Fukushima
Dai-ichi
nuclear power plants disaster
on 11-March-2011 in Japan
caused huge damage to the
environment not only onsite
and nationally but also remote
and world wide. Contamination
of
131
I and
137
Cs were found in
agricultural, animal and marine
products
which
resulting
negative effects on producers Figure 1.2 Dependence of the integrated activity
concentration of radionuclides during March
and consumers (Fujiwara et al., and April 2011 on the distance from Fukushima
2012). Maps covering more
than 2000 locations within 100 km from the Fukushima Dai-ichi nuclear power plants
contaminated with Cs was established in Japan (MEXT, 2011a). Furthermore,
contamination was also found in the nearby countries such as Korea (Kim et al., 2012),
and Vietnam (Long et al., 2012) (Figure 1.2 and 1.3).
10
Figure 1.3 Comparison between Third Airbone Monitoring Results and Map of Cs-134
Concentration
(source: MEXT (Ministry of education, culture, sports, science and technology, Japan, 2011a)
1.2.4 The Chemical behavior of Cs in soils
In soil, the mobility of Cs is very low (from 0.11 to 0.29 cm/ year (Schuller et al.,
1997). It does not exist at a depth below 40 cm and is mostly present in the 20 cm soil
layer (Korobova et al., 1998, Takenaka et al., 1998). When present in the soil, the fate
of Cs will be one of the following: (1) absorbed by soil particles (mostly clay), (2)
mobile with soil solution, and (3) accumulated by plants and microorganisms (Bunzl et
al., 1989, Beckmann and Faas, 1992, Bunzl et al., 1994).
Clays and potassium rich soils retain Cs in the interlayer positions through cation
exchange (Paasikallio, 1999). Because Cs has low hydration energy, it can be absorbed
by clay particles. As a result, it is difficult for plants to take up Cs (LaBrecque and
Rosales, 1996, WHO, 1983b). The Cs adsorption does not correlate with soil pH,
water content, cation exchange capacity, and exchangeable calcium (Paasikallio,
1999).
11
1.2.5 Caesium and health effect in humans
Stable Cs exposure can be low risk because of low levels in the environment. Studies
on the levels of stable Cs in humans are limited. It was detected in the urine of U.S
citizens at a range of 4-5 μg/L which varies with age and gender. There were no
reports on death, systemic, immunological and lympho reticular, neurological,
reproductive, development and cancer effects through inhalation of stable Cs.
Exposure to stable Cs through the oral pathway or skin is of very low risk to humans
(cited in ATSDR, 2004) (ATSDR, 2004).
134
Cs and
137
Cs produce gamma radiation
which is a health hazard. Following overexposure to radiocaesium, gamma rays can
damage tissues and internal organs. (Bartstra et al., 1998, Matsuda et al., 1985, Nikula
et al., 1995, Nikula et al., 1996, Padovani et al., 1993, Ramaiya et al., 1994, Skandalis
et al., 1997).
1.2.6 Caesium and environmental effects
There are few studies on stable Cs and environmental effects. Studies on laboratory
animals revealed that stable Cs is of low toxicity. For rats and mice the oral LD 50 was
from 800 to 2,000 mg/kg Cs and Cs iodide or chloride is less toxic than Cs hydroxide.
In female mice, an oral dose from 125 to 500 mg/kg CsCl can increase chromosomal
breaks in bone marrow cells (Ghosh et al. 1990, 1991 cited in ATSDR, 2004).
However, radiocaesium has been detected in bird, animal, fish, and plants, proving that
Cs can be bioaccumulated. Studies revealed that there was a transfer of
to reindeer and the concentration of
137
137
Cs from soil
Cs did not decrease during the study period in
reindeer (Skuterud et al., 2005). Waston et al. discovered radiocaesium contamination
in marine mammals (seal and porpoises). They found that the level of radiocaesium
was correlated with body weight and 100 times higher compared to sea water. They
also discovered that the concentration of radiocaesium decreased with distance. The
study indicated that in muscle the
137
Cs level was higher than in liver (Watson et al.,
1999).
In terrestrial organisms,
137
Cs was equally found in the legs and head of Geotrupine
beetles (Cleoptera: Scarabaeoidea) in areas of Poland after the Chernobyl accident
(Mietelski et al., 2003).
12
1.2.7 Remediation of Cs polluted soils
After the Fukusima- Dai ichi nuclear power plant disaster, the Japan Atomic Energy
Agency (JAEA) has suggested some solutions to decontaminate radiocaesium
pollution. In forests, removal of fallen leaves and humus, topsoil stripping, tree trunk
washing, branch trimming and felling were applied. In farmland, reversal tillage and
topsoil stripping (with and without solidifying agent usage) were used. In houses and
large buildings, high-pressure water jet-based washing incorporating with nanobubbles or 3% hydrogen peroxide or ozone water and steel brush were
employed(Japan Atomic Energy Agency (JAEA), 2012).
1.3 Lead
1.3.1Characteristics of Pb
Lead is a metal that exists in the Earth’s crust and ranges in concentration in surface
soils worldwide from 10 to 67mgkg-1. Generally, it is present as compounds with other
elements such as sulphur in PbS and PbSO4 or oxygen in PbCO3. In nature, there are
four stable isotopes of Pb which are
208
Pb (51-53%),
206
Pb (23.5-27%),
207
Pb (20.5-
23%), and 204Pb (1.35-1.5%).
Lead is produced less than Fe, Cu, Al and Zn. In the U.S, about half of Pb is used for
battery manufacture. Other Pb applications are soldering, bearings, cable cover,
ammunition, plumbing, pigments, and caulking. Some other metals used with Pb are
antimony (Sb), calcium (Ca), tin (Sn), silver (Ag), strontium (Sr), tellurium (Te) in
maintenance-free storage batteries, anodes, electroplating processes, chemical
installations and nuclear shielding, sleeve bearings, printing, and high-detail castings
(Manahan, 2003).
Forms of Pb generally released into soil, groundwater, and surface waters are ionic Pb,
Pb(II), Pb oxides and hydroxides, and Pb metal oxyanion complexes. Among them
Pb(II) and lead-hydroxy complexes are the most stable forms. Pb(II) is well-known as
an reactive form, forming mononuclear and polynuclear oxides and hydroxides such as
Pb(OH)2 which is the most stable solid in the soil. PbS is formed in the increasing
presence of sulfide and in reducing conditions. Tetramethyl Pb is an organolead
formed under anaerobic conditions and with microbial alkylation (GWRTAC, 1997).
13
Lead phosphates, and Pb carbonates are formed when the pH is greater than 6, and Pb
(hydro) oxides are predominant insoluble forms (Raskin and Ensley, 2000).
Lead compounds are mainly ionic (e.g., Pb2+SO42-) and covalent (e.g., tetraethyl Pb
Pb(C2H5)4), some Pb(IV) compounds are strong oxidants such as PbO2. Lead
Pb(OH)2.2PbCO3 is one of the salts that is formed by Pb (II) which was mostly used in
white paint and chronically poisoned children who consumed white paint peel. Among
many useful compounds of Pb (II) and (IV), Pb dioxide and sulphate are most
common. They participate in the charge and discharge of Pb storage batteries. Lead
also exists in many organic forms as organo-Pb such as tetraethyl Pb. This is well
known because of its application in gasoline additive. There are more than 1000
organolead compounds which are methyl, ethyl and their salts (dimethyldiethyllead,
trimethyllead, and diethyllead dichloride) (Wuana and Okieimen, 2011)
1.3.2 Sources of Pb
In the atmosphere, Pb occurs primarily as PbSO4, and PbCO3. Non-organic Pb
compounds can exist in the particulate form (U.S. ATSDR, 2005). According to a
study from 1977, the average particle size of Pb emitted from smelters was 1.5 μm in
which 86% of the particle sizes were smaller than 10 μm (Corrin and Natusch, 1977),
as cited by U.S. ATSDR, 2005). The particle size of Pb emissions may vary. It is
associated with the high-temperature combustion processes. The smallest size was < 1
μm (U.S. ATSDR, 2005).
In the aquatic environment, Pb can be found in highly mobile and bioavailable ionic,
organic complexes with dissolved humus materials which bind strongly and limits
availability, colloid particles (iron oxide) which binding strongly with low mobility, or
clay particles, or dead remains organisms (very low mobility and availability) (OECD,
1993). The form of Pb is affected by pH, salinity, sorption and biotransformation
processes. In acidic environments, Pb present as PbSO4, PbCl4, ionic, Pb hydroxide
Pb(OH)2 (U.S. EPA, 1979).
14
In water, tetraethyl-lead and tetramethyl-lead is easy to photolyse and volatilize.
Trialkyl compounds (carbonates, hydroxides, and halides) then degraded to dialkyl and
finally to inorganic Pb oxides. Some of these degradations are more difficult than the
original tetraalkyl-lead compounds (DeJonghe et al., 1981), as cited by U.S. ATSDR,
2005). Lead in surface water is derived from (1) biogenic material, (2) Aeolian
particles, (3) fluvial particles and (4) erosion (Ritson et al., 1994), as cited by U.S
EPA, 2005a). In the open ocean, 90% of Pb is in the dissolve phase (Reuer and Weiss,
2002). 50-70 percent of this Pb are organic ligand complexes (Reuer and Weiss, 2002,
as cited by U.S. EPA, 2005a). Specifically, Pb in surface water can reside less than 5
years (Gobeil et al., 2001), as cited by (Macdonald et al., 2005).
Groundwater contains very low levels of Pb due to the high binding of Pb to soil
particles and humus, and the diffusion of Pb from deposits is a relatively low process
(Hansen et al., 2004a). The mobility of Pb in soil depends on soil pH and organic
content. The sorption and immobility of Pb decreases the bioavailability to humans
and terrestrial organisms (OECD, 1993).
1.3.3 Lead in soils
Lead deposits in soils from anthropogenic sources such as mining and smelting
activities, manures and sewage sludge application in agriculture and vehicle exhausts
contamination. Pb has been applied to orchard trees as Pb arsenate (PbHAsO4) to
control insect pests, therefore elevated concentration of Pb are found in orchard soils
(Frank et al., 1976, Merry et al., 1983).
The general concentration of Pb in uncontaminated soils worldwide is 17 mg/kg
(Nriagu, 1978). During the early decades of the last century, due to the development of
the internal combustion engine, there was a growing demand of Pb for petrol of higher
octane ratings to avoid uneven combustion in the engine cylinders. Pb alkyls
(tetraethyl and tetramethyllead) were discovered in the early 1920s as an addition that
helped over come the problem. In 1923, the first leaded petrol was sold and rapidly
became standard.
15
Warren and Delavault, 1960 reported that soils and vegetation near the roads contained
unusual high level of Pb and petrol fume Pb had special attention (Warren and
Delavault, 1960). Cannon and Bowles (Cannon and Bowles, 1962) found that Pb was
present in grass within 152 m downwind of road in Denver, Colorado, USA and there
was a relationship between Pb content and distance. There was a zone of
approximately 15 m on both side of a large number of roads that were contaminated
with high levels of Pb compared with local background and it was evident that there
was worldwide contamination due to the use of leaded petrol.
Lead can be carried over long distance in the form of vehicles exhausts or industrial
fumes (Nriagu and Pacyna, 1988). A decrease of an average of <120 mg Pb/kg in the
south of Norway to <10 mg Pb/kg in the far north was observed by Steinnes (Steinnes,
1984) and the lower values of the remote areas caused by western European industries.
Generally, atmospheric deposition of Pb was from 3.1 to 31 mg/m2/year in distant and
countryside locations and from 27 to 140 mg/m2/year in industrial and suburban
locations, 0.4 g/ha/year at the South Pole, 7.2 g/ha/year in north west Canada and 6.3
g/ha/year in northern Michigan, USA (Sposito and Page, 1984). Due to the fact that
many countries have reduced or even phased out the use of Pb in petrol (Nriagu,
1990d), the Pb content in soils has decreased (Jones et al., 1991, Jones and Johnston,
1991).
1.3.4 The chemical behavior of Pb in soil
Lead is present primarily in the top soil layers and is difficult to leach. The soils tend
to keep Pb immobile with their organic matter (Lounamaa, 1956, Zimdahl and
Skogerboe, 1977, as cited in (Alloway, 1995)). Similar to other metals Pb remains
insoluble in lead-organic matter complexes (Steinnes, 1984). The most common
existence of Pb in soil is the +2 oxidation state for example insoluble PbS is generated
in reduced soil. In the oxidative condition Pb occurs as Pb+2 ions, and is less soluble as
a result of producing complexes with organic matter, binding with oxide and silicate
clay minerals, precipitating as carbonate, sulfate, or phosphate. In pH > 7 soils, Pb
solubility may increase because of the formation of Pb-organic and Pb-hydroxy
complexes (McBride, 1994). Pb activities depend much on organic matter in soil. At
pH ≥ 4 Pb binds strongly to humic matter (Bunzl et al., 1976, Kerndorff and Schnitzer,
16
1980). Pb can bind to clays, silts, iron and manganese oxides (U.S. ATSDR, 2005).
Sipos et al. claimed that clay minerals absorb less Pb than organic matter (Sipos et al.,
2005). Reduction of Mn and Fe oxides and higher pH may induce Pb release into soil
solution from the solid phase (Charlatchka and Cambier, 2000). Hansen et al. stated
that there are very small amount of Pb in the soil solution. Acidic conditions in soil
can increase Pb mobility and bioavailability (Hansen et al., 2004a)
1.3.5 Lead and health effects in humans
Depending on the level, duration and timing, exposure to Pb can cause a range of
biological effects. Pb is a toxicant to multiple organ systems. The effects may vary
from enzyme inhibition and anemia to nervous, immune and reproductive system
disorders, kidney and cardiovascular impaired functions, and death. Children are more
sensitive than adults because of their rapid growth and maturation (IPCS, 1995).
Lead is well known as a neuro-toxicant. Abnormal neurodevelopment in children is
one of the most common effects due to exposure in uterus and early childhood. The
lifetime exposure of the mother is very important since Pb accumulates in the skeleton
and is mobile during the pregnancy. Lactation can harm fetuses and breastfed children
(WHO/UNECE, 2007). Lead exposed children have been shown to have low
intelligent quotients (IQ), behavior deficiency, and learning abilities. Lead exposure at
a level of 10-15 μg/dL results in such the effects. However, the effects do occur at Pb
level less than 10 μg/dL and there maybe no threshold (Canfield et al., 2003). Some
reports have shown the effects at blood Pb levels less than 5 μg/dL, however there are
less cases at this level (U.S. EPA, 2006). At levels of exposure above70 μg/dL in
blood, children can have severe neurological effects, lethargy, convulsions, coma and
death. Adult nervous systems can be affected. Evidence of long term exposure to Pb
include decreased performance, finger, wrist, and ankle weakness (U.S. ATSDR,
2005).
1.3.6 Lead and environmental effects
Lead affects birds through ingestion of Pb shot and fishing sinkers. It causes adverse
effects on Common loon, Trumpeter, Mute and Tundra swans, Sandhill cranes, and
other bird of similar feeding habits or those living in contaminated sites. In the
17
digestive system, Pb is slowly ground down, this results in a high level of Pb in blood,
kidney, liver and bone (IPCS, 1989). In water birds, Pb in blood of 0.5 mg/kg is
considered toxic. Death is observed at 20-40 mg/kg, sub-lethal symptoms at 7.5
mg/kg. The lethal level is 5.0 mg/kg or more (10-14 μg/g dry weight) in the liver. In
experiments, the lethal level of Pb ranges from 5-80 mg/kg (U.S. EPA, 1994).
In wild animals, there are few reports on Pb toxicity. In all laboratory species, Pb has
fatal effects in some organs and organ systems which are circulatory system, central
nervous system, and kidney, reproductive and immune system (IPCS, 1995).
With terrestrial organisms, a Pb level of 50-60 mg/kg dry weight of soil has been
suggested by the CSTEE (EU’s scientific committee on environmental toxicology) as
the predicted No-Effect Concentration (PNEC) (Tukker et al., 2001). In nematodes,
impaired reproduction was observed if they had ingested Pb contaminated bacteria and
fungi. In microorganisms, Pb shows effects at 10 mg/kg soil. Although Pb prevents
nitrogen mineralization, Pb compounds do not seem to be toxic to microorganisms.
Moreover, inorganic compounds of Pb are less toxic than organic such as trialky and
tetraalkyl. Microorganisms can tolerate and increase their tolerance to Pb (IPCS, 1989,
(Bieby Voijant Tangahu et al., 2011).
In plants, Pb shows its effects only at a very high concentration. Pb is taken up by
plant roots with limited translocation to shoots. Pb strongly binds to soils and
inorganic Pb forms insoluble salts and complexes. This results in low availability to
plants. However, Pb can become available in an acidic medium. The level to cause
toxicity effects on plants (phytosynthesis, growth ect.) can be from 100 to 1000 mg/kg
(IPCS, 1989). Moreover, in plants, most of the Pb is located on the leaves as a surface
coating not deposited into the plant (OECD, 1993).
1.3.7 Remediation of Pb polluted soils
Remarkable efforts have been made over the past 20 years to reduce the toxicity of Pb
in contaminated soils. The solutions suggested were either in situ stabilization or
removal by chemical and biological methods. Pb contaminated agricultural soils seems
to be favorable for in situ stabilization with phosphate amendments (Mulligan et al.,
18
2001). Soil washing (Dermont et al., 2008), and phytoextraction were also suggested.
Hyperaccumulating plants were used to remediate high level of Pb in soil (Huang and
Cunningham, 1996). Organic agents were used to enhance Pb accumulation by plants
(Huang et al., 1997, Shen et al., 2002).
1.4 Remediation of contaminated soils
The purpose of remediation is to find out the best solution that reduces the
concentration of contaminants or their availability to a level that is safe for human
health and the environment. For heavy metals in soils the physical and chemical
properties of the metals strongly influence the choice of appropriate treatment methods
(Martin and Ruby, 2004). In addition, the information on the site such as the physical
characteristics, and the type, level and distribution of contaminants and an overview of
the situation is measured to propose remediation solutions. Then, the desired level of
metals will be determined based on the soil quality standards or site-specific risk
assessment. The objective of remediation may focus on the total concentration or
leachable metals or both (Martin and Ruby, 2004).
There are different classifications of remediation methods. Gupta et al. categorized
methods based on hazard-technologies: (1) gentle in situ remediation, (2) in situ harsh
soil restrictive measures, and (3) in situ or ex situ harsh soil destructive measures
(Gupta et al., 2000). The USEPA classified methods into (1) source control (in situ and
ex situ) and (2) containment remedies (vertical engineered barriers (VEB), caps, and
liner) (USEPA, 2007). Another classification divides remediation into five categories:
isolation, immobilization, toxicity reduction, physical separation, and extraction
(GWRTAC, 1997). The factors that affect the selection of these remediation measures
are: (1) cost, (2) long-term application, (3) commercial potential, (4) general
acceptance, (5) high concentration applicability, (6) mixed wastes ability (containing
organic wastes), (7) toxicity decrease, (8) mobility reduction and (9) mass reduction
(Wuana and Okieimen, 2011). In summary, remediation methods fall into three
categories: Physical techniques (using containment or removal), chemical techniques
(that either increase or decrease the mobility of contaminants then extract or reduce the
exposure potential), and biological techniques (use natural or biochemical processes to
19
either elevate mobility and then release metals for extraction or reduce the availability
(Knox et al., 2001).
1.4.1 Remediation of heavy metals in soils
At present, immobilization (solidification/ stabilization and vitrification), soil washing,
and phytoremediation have been used to remediate heavy metal contaminated soils
(Wuana and Okieimen, 2011). Immobilization techniques can be applied in situ or ex
situ. Application ex situ is when highly contaminated soils occur. The soil must be
removed from the original place so that it cannot harm the population and ecosystem.
Fast and easy application and relative low cost of investment and operation are
advantages of this method. Conversely, it can cause high invasivity, create more solid
wastes, raise byproduct storage issues, release more harm if surrounding
physicochemical properties of the soil change, and end up with permanent operation
cost. For the in situ technique, the contaminated soils do not need to be excavated or
removed from original areas but the fixing amendments are introduced. The
amendments reduce the mobility and toxicity of heavy metals in soils. The advantages
of this are this method is less invasive, simple and rapid, quite cheap, produces limited
wastes, has highly acceptance, and widely covers many inorganic pollutants. By
contrast, this technique is only temporary, and surface application pollutants can be
reactivated, and permanent monitoring is required (Martin and Ruby, 2004, Gupta et
al., 2000, USEPA, 2007, 1997).
In immobilization technology, organic and inorganic amendments (Table 1.2 and 1.3)
which are clay, cement, zeolites, minerals, phosphates, organic composts, and
microbes are often used (GWRTAC, 1997, Finžgar et al., 2006). Recently, industrial
residues such as red mud (Boisson et al., 1999, Lombi et al., 2002, Anoduadi et al.,
2009) and termitaria (Anoduadi et al., 2009) have been considered to be low cost
sources. The major function of these amendments is to make heavy metals more stable
through sorption, precipitation, and complexation (Hashimoto et al., 2009). However,
the exact mechanism of this process has not been figured out due to the complicated
matrix in soil and current analytical techniques (Wang et al., 2009).
20
Solidification often uses binding agents to keep contaminants in a solid product and
reduce the external accessibility via a combination of chemical reactions,
encapsulation, and permeability or surface area reduction. Stabilization (fixation) uses
reagents that react with heavy metals in contaminated soils and result in producing
more chemically stable constituents (Evanko and Dzombak, 1997).
Table 1.2 Organic amendments for heavy metal immobilization
(Guo et al., 2006).
Material
Heavy metal immobilized
Bark saw dust (from timber industry)
Cd, Pb, Hg, Cu
Xylogen (from paper mill wastewater)
Zn, Pb, Hg
Chitosan (from crab meat canning industry)
Cd, Cr, Hg
Bagasse (from sugar cane)
Pb
Poultry manure (from poultry farm)
Cu, Pb, Zn, Cd
Cattle manure (from cattle farm)
Cd
Rice hulls (from rice processing)
Cd, Cr, Pb
Sewage sludge
Cd
Leaves
Cr, Cd
Straw
Cd, Cr, Pb
Table 1.3 Inorganic amendments for heavy metal immobilization
(Guo et al., 2006).
Material
Heavy metal immobilized
Lime (from lime factory)
Cd, Cu, Ni, Pb, Zn
Phosphate salt (from fertilizer plant)
Pb, Zn, Cu, Cd
Hydroxyapatite (from phosphorite)
Zn, Pb, Cu, Cd
Fly ash (from thermal power plant)
Cd, Pb, Cu, Zn, Cr
Slag (from thermal power plant)
Cd, Pb, Zn, Cr
Ca-montmorillonite (mineral)
Zn, Pb
Portland cement (from cement plant)
Cr, Cu, Zn, Pb
Bentonite
Pb
Vitrification treatment uses high temperatures on the contaminated sites to create
vitreous material (usually oxide solid). This can also remove organic contaminants by
destroying and/ or volatilization. This technique can be applied to reclaim
contaminated soil with heavy metals such as Pb, Cd, Cr, asbestos, and asbestos
containing materials. Volatile metals must be collected for treatment or disposal.
21
Vitrification can be applied in situ and ex situ with a large range of organic and
inorganic contaminants. In situ is preference because of lower cost and energy
requirement (USEPA, 1992).
Ex situ treatment includes excavation, pretreatment, mixing, feeding, melting, heating,
output gas collection and treatment, and casting of melted product formation. Energy
for the melting plays a major role in cost. The waste from the process can be recycled
and reused (Smith et al., 1995). In situ applications can also use a current to create heat
to melt contaminated soils. A single set with four electrodes can treat up to 1000 tons
with a depth of 20 feet contaminated soil (3 to 6 tons per hour). If the soil is too
alkaline (containing too high Na2O and K2O) or too dry, supplement such as flaked
graphite and glass grit must be added (Buelt and Thompson, 1992).
Soil washing is a volume reduction or waste minimization technique. It can be
conducted in situ or ex situ. First of all, the contaminants containing soil particles are
separated from bulk soil. Contaminants are separated from the soil and recovered by
chemicals. Acids and chelators are commonly added (Dermont et al., 2008). Heavy
metals are insoluble and absorbed in soil. The addition of acid, alkalis, complexes,
solvents, and surfactants enhances the removability of heavy metal from the soil into
the aqueous phase. The contaminants removed may be further treated (by chemical
sorption on activated carbon or ion exchange) (Dermont et al., 2008), thermal, or
biological procedure) or go to hazardous waste landfill. The remaining soil can be
recycled, filled in another area, or disposed of as non-hazardous waste.
The efficiency of the soil washing method can be >70-80%. This solution is cost
effective with sandy and granular soils but not for clay and silt (less than 30-35%). The
strong binding between heavy metals and soil especially clay makes the treatment
difficult. Therefore, extractants that have a high potential for dissolving heavy metals
and preserving the properties of the soils must be considered (Tejowulan and
Hendershot, 1998). In this case organic acid and chelating agents are suggested (Ju and
Klarup, 1994). Oxalic, citric, formic, acetic, malic, succinic, malonic, maleic, lactic,
aconitic, and fumaric acids are natural low molecular weight organic acids
22
(LMWOAs) used to dissolve heavy metals. Some chelators such as citric acid (Naidu
and Harter, 1998), tartaric acid (Ke et al., 2006), and EDTA (Peters, 1999, Tejowulan
and Hendershot, 1998, Sun et al., 2001) are not proven for full-scale application.
1.4.2 Remediation of radionuclides in soils
Currently used techniques to remediate radionuclide contaminated soils include
physical approaches, agriculture-based countermeasures, and phytoremediation (Zhu
and Shaw, 2000). Physical approaches may require the removal of contaminated soil to
treat with a range of dispersing and chelating agents (Entry et al., 1996). This method
is not applicable on a large scale due to the labor, equipment, and cost demand. In
addition, it can spread the contaminant while transporting and disturbing the
ecosystem. The use of dispersing and chelating chemicals may cause harm to the
environment (Entry et al., 1996). Some solutions that can be used to reduce the
impacts of contamination are spraying with detergents or cleaning agents, harvesting
and removing plant portions, soil ploughing and scrap and removing surface soil (Zhu
and Shaw, 2000).
Agriculture-based countermeasures have been suggested as one of the most effective
ways to reduce to the level of radiation to humans through crops by adding mineral
and chemical absorbents or K- and Ca-containing fertilizers. Natural or synthetic
amendments reduce the phytoavailability of radionuclides (Zhu and Shaw, 2000). An
example of natural or artificial zeolite increases the Kd (soil-liquid distribution
coefficient) of radiocaesium. Another chemical is ammonium-ferric-hexacyano-ferrate
(II) (AFCF). At a level of 10 and 100 g/cm2 AFCF can reduce the radiocaesium from
sandy soil to ryegrass without growing impact on plant (Vandenhove et al., 1996). It
was observed that the K- and Ca-containing fertilizer application can reduce the uptake
of radiocaesium and radiostrontium respectively (Jackson et al., 1965, Evans and
Dekker, 1968).
Bioremediation is considered an energy saving and cost effective solution which
includes soil fungi, mycorrhizae and phytoextraction application (Zhu and Shaw,
2000). Stable Cs accumulation was found in 18 fungal species (Clint et al., 1991).
Radioceasium immobilization by soil fungi was discovered by Dighton et al. (1991).
23
This can be used to block the radionuclide in the top soil layer and reduce groundwater
pollution (Dighton et al., 1991). Mycorrhizae are an important group of soil fungi. The
role of mycorrhizae in radionuclide contaminated soil was investigated and needs to be
further studied (Zhu and Shaw, 2000).
1.4.3 Phytoremediation of heavy metals and radionuclides
Phytoremediation
also
known
as
green
remediation,
botanoremediation,
agroremediation, or vegetative remediation- “an in situ remediation strategy that uses
vegetation and associated microbiota, soil amendments, and agronomic techniques to
remove, contain, or render environmental contaminants harmless” (Cunningham and
Ow, 1996, Helmisaari et al., 2007). The concept has been used for 300 years in
wastewater treatment, but the idea of using plants to accumulate heavy metals was
proposed in 1983 (Chaney et al., 1997, Henry, 2000). Using terrestrial plants to
accumulate heavy metals is a new technology. However, it has been used for full-scale
remediation in many countries in the world.
Phytoremediation can be used to decontaminate not only heavy metals (e.g. Ag, Cd,
Co, Cr, Cu, Hg, Mn, Mo, Ni, Pb, and Zn) but also radionuclides (e.g. 3H, 90Sr,
239
137
Cs,
Pu, 234U, 238U) in soils (Andrrade and Mahler, 2002, Negri and Hinchman, 2000). In
soils, a radionuclide reacts the same ways as a nonradioactive isotope. Hence, its
physical concentration is lower than the nonradioactive isotope (Wild, 1993).
The advantages of this method are (1) it does not require expensive or high technique
equipment and it can share agricultural resources, (2) does not disrupt the environment
and prevents erosion, (3) does not need to wait for new colonization by plants, (4) does
not need disposal sites, (5) accepted by the public due to being aesthetically pleasing,
(6) does not need excavation and transportation of contaminated soils, (7) can deal
with multi pollutants sites (organic or inorganic (Table 1.4)) , (8) energy saving, (9)
has a wide range and is a permanent application of low cost (Raskin and Ensley,
2000).
The disadvantages are (1) efficacy depends plant growth condition (i.e., temperature,
humidity, soil condition, rain volume, diseases), (2) requires agricultural knowledge
24
and equipment when applied on a large scale, (3) contaminated tissues may be go back
to the environment during autumn or be collected to be used as fuel, (4) may be
consumed by pets, birds and wild animals, (5) requires a long time for treatment.
Phytoremediation
technologies
include
phytoextraction
(phytoaccumulation),
phytostabilization, phytofiltration (rhizofiltration), phytovolatilization (Dushenkov,
2003).
Table 1.4 Substances amenable to the phytoremediation process
Organics
Inorganics
Chlorinated solvents
Metals
TCE, PCE, MTBE, carbon, tetrachloride
B, Cd, Co, Cr, Cu, Hg, Ni, Pb, Zn
Explosives
Radionuclides
TNT, DNT, RDX, and other nitroaromatics
137
Cs, 3H, 90Sr, 235,238U, 95Nb, 99Tc, 106Ru, 144Ce,
226,228
Ra, 239,240Pu, 241Am, 228,230,232Th, 244Cm,
237
Np
Pesticides
Atrazine,
Others
bentazon,
and
other
chlorinated
and
As, Na, NO3, NH4, PO4, perchlorate (ClO4)
nitroaromatic chemicals
Wood preserving chemicals
PCP and other PAH’s
Source: (Glass Associates, 1999, Dushenkov, 2003)
Phytoextration (phytoaccumulation) is “the uptake/ absorption and translocation of
contaminants (heavy metals/ radionuclides) by plant roots into the above ground
portions of the plants (shoots) that can be harvested and burned gaining energy and
recycling the metal from the ash” (Tangahu et al., 2011, Dushenkov, 2003).
Heavy metals will be absorbed at the root surface then enter the root cells through
cellular membrane. A proportion of the metals will be immobilized in the vacuole of
roots. Some will be intracellularly transferred by the root xylem to the stems and
leaves (Lasat, 2000). Most metals exist as insoluble forms when inside the plant such
as in carbonates, sulphates, or phosphates (Raskin and Ensley, 2000). Phytoextration
has two trends: one which is continuous or natural phytoextraction and the other
chemically enhanced phytoextraction (Lombi et al., 2001, Ghosh and Singh, 2005).
25
Continuous or natural phytoextraction relates to the use of natural hyperaccumulator
plants which have high accumulation potential. These plants can accumulate more than
10 mgkg-1 Hg, 100 mgkg-1 Cd, 1000 mgkg-1 Co, Cr, Cu, and Pb; 10000 mgkg-1Zn and
Ni (Baker and Brooks, 1989, Lasat, 2002). An estimation of 400 plant species from 45
families have been reported to be hyperaccumulators such as Brassicaceae, Fabaceae,
Euphorbiaceae, Asterraceae, Lamiaceae, and Scrophulariaceae (Salt et al., 1998,
Dushenkov, 2003). Thlaspicaerulescens, Ipomea alpine, Haumaniastrumrobertii,
Astragalusracemosus, Sebertia acuminate can accumulate very high levels of Cd/ Zn,
Cu, Co, Se, and Ni, respectively (Lasat, 2000). Salix viminalisL. , Brassica juncea, L.
Zea mays L., and Helianthus annuus L. can uptake and tolerate high levels of heavy
metals (Schmidt, 2003a). However, the disadvantages of using hyperaccumulating
plants relate to their slow growth and low biomass. Therefore, chelator amendments
have been suggested to enhance the bioavailability of heavy metals in soil (Nowack et
al., 2006).
Chelate-Assisted (Induced) phytoremediation. In the last decade, the applications of
chelating agents to enhance phytoextraction have received much attention due to the
cost effective and positive effects in increasing heavy metal solubility. The synthetic
or natural organic ligands such as diethylenetriaminepenta acetic acid (DTPA),
glycoletherdiaminetetraacetic acid (EDGA), ethylene diaminetetraacetic acid (EDTA),
ethylene diaminedisuccinate (EDDS), nitrilotriacetic acid (NTA), low molecular
weight organic acids (LMWOAs) and humic substances (HSs) have high affinity with
metals (Ferrand et al., 2006, Quenea et al., 2009, Yip et al., 2010). These agents
remove heavy metals from the soil matrix into soil solution by forming metal-chelate
complexes. The complexes are taken up through a positive apoplastic pathway by
plant roots and translocated to above ground portions. However, heavy metal leaching
is the concern when using organic amendments. The plants can only take up a small
part of the extracted heavy metals in soil solution (Evangelou et al., 2007).
Humicacid (HA) has been used as an organic chelator to enhance the bioavailability of
Pb and Cs. HA is soluble at pH>4 (Stevenson, 1994). Humic acids have carboxyl (COOH) and phenolic (-OH) groups (Steinbüchel and Hofrichter, 2001) that function as
26
organic macromolecules in transport, bioavailability, and solubility of metals (Lagier
et al., 2000). Researchers have showed that HA increased desorption of Cd, Mo, Pb,
and B and enhanced the accumulation of these elements in Zea mays L. and
sunflowers (Helianthus annuus L.). Because of a lower constant stability (compared to
synthetic agents), HA is a good option for phytoextraction and prevents the leaching of
heavy metal-humic acid complexes through the soil profile (Chen and Aviad, 1990,
Mackowiak et al., 2001, Zhang et al., 2003).
Efficiency of phytoextraction
Depending on the level of contamination and the target concentration of remediation,
phytoextraction may be applied for several seasons to reduce contamination to the
acceptable level. Generally, the bioaccumulation factor (BF) is used to estimate the
efficacy of phytoextraction of metals and radionuclides. This is also known as the
transfer factor (TF) (Cook et al., 2007, Dushenkov, 2003). Extracted metals
(M)(mg/kg), and treatment duration (tp) are also used. In these equations, n is the
cropping times in one year of that chosen crop, assuming that the heavy metal
contamination occurs in the top layer (0-20 cm), bulk density is 1.3 t/m3, equal to
26,000 t/ha of soil mass ((Zhuang et al., 2005) cited in (Wuana and Okieimen, 2011))
BF/TF=
M (mg/kg plant) =Metal concentration in plant tissue
Biomass
tp (year) =
Phytostabilisation is “the use of certain plant species to immobilize the contaminants
in the soil and groundwater through absorption and accumulation in plant tissues,
adsorption onto roots, or precipitation within the root zone preventing their migration
in soil, as well as their movement by erosion and deflation” (Tangahu et al., 2011).
27
Phytofiltration (Rhizofiltration) is “the adsorption or precipitation onto plant roots or
adsorption into and sequesterization in the roots of contaminants that are in solution
surrounding the root zone by constructed wetland for cleaning up communal
wastewater (Tangahu et al., 2011).
Phytovolatilizationis “the uptake and transpiration of a contaminant by a plant, with
release of the contaminant or a modified form of the contaminant to the atmosphere
from the plant” (Tangahu et al., 2011).
1.5 Sunflowers for phytoremediation of trace metals in soils
There is a wealth of evidence that several species of plants can tolerate heavy metal
and other chemicals for example Indian mustard (Brassica juncea), Corn (Zea mays
L.) or sunflower (Helianthus annuus L.).Therefore these plants can be used for
phytoremediation.
Sunflower (Helianthus annuus L.) (Figure 1.4) known as an oil yielding and
ornamental plant is one of the most important environmental clean up crops. It can be
applied in chemical and radiological contaminated soil and water. Within one hour of
an experiment, the concentration of Cd, Cr, Cu, Mn, Ni, and Pb in soils were reduced
significantly with sunflowers (Eapen et al., 2007). Sunflower incorporated with Indian
mustard (Brassica rapa) was applied to remediate the USA heavy metal and
radionuclide contaminated sites for example, Pb in Connecticut from 1997-2000 and
Pb contaminated soil (75-3,450 mg/kg) at the Daimler Chysler car manufacturing
company. In one season, with sunflower and Indian Mustard application, the level of
Pb was reduced to 900 mg/kg. Sunflowers were also used to clean up uranium (U) (47
mg/kg of soil) at US Army sites at Aberdeen, Maryland. It can accumulate uranium
from 764-1669 mg/kg. Other researchers also demonstrated that sunflowers are able to
accumulate uranium (U) (Salt et al., 1998 and Jovanovia et al., 2001). The cost of
sunflower application (combine with Indian Mustard) was 40-50$ US per cubic yard.
In comparison with traditional excavation and landfill disposal it saves more than US$
1.1 million in the USA(Singh et al., 2007). Sunflowers were found to be efficient in
decontamination heavy metals and radionuclides by Dushenkov et al. (1995), Salt et
al. (1998), and Jovanovia et al. (2001). Sunflower roots can reduce the level of Cd, Cr
28
(VI), Cu, Mn, Ni, Pb, Sr, U (VI), and Zn to discharge limits in water after 24 hours
(Dushenkov et al., 1995). Moreover, sunflowers were considered as a potential
candidate for phytoremediation, because of their high biomass production. This
contributes to increasing the amount of contaminants removed from polluted soils. In
general, hyperaccumulators often have slow growth and produce low biomass.
Therefore sunflowers are suitable for heavy metal contaminated remediation of soil
with fast growth, high biomass, and high tolerance and accumulation of metals and
radionuclides (Pilon-Smits, 2005)(all cited in (Angelova et al., 2012).
Figure 1.4 Sunflowers (Helianthus annuus)
Source: http://homeopathtyler.wordpress.com
1.6 Earthworms for remediation of heavy metals and radionuclides.
Using
earthworms
(vermiremediation
technology)
for
soil
heavy
metal
decontamination is also an innovative and effective solution. According to Hand
(1988) (Hand, 1988) this technology is easy to perform. Vermiremediation of soils
contaminated with chemicals would cost less than two times ($500-1000 per hectare)
as compared with mechanical excavation ($10,000 - 15,000 per hectare). This is due to
the fact that: (1) population of earthworms increases significantly. It takes only 3
29
months to create a population of 0.2 - 1.0 million earthworms in one hectare of
polluted land (Sinha et al., 2008). (2) Earthworms have a high tolerance and
accumulate a variety of toxicants from soils such as heavy metals and organic
substances (Ireland, 1979). The ability of reducing heavy metals in the biosolids by
earthworms was documented by Contreras-Ramos et al. (2006) (Contreras-Ramos et
al., 2006). (3) Earthworms also have the ability to inhibit pathogens in biosolids such
as Salmonella (Shakir Hanna and Weaver, 2002). This is important in biosolids
treatment since Salmonella is one of the most important quality parameters to assess
biosolids. (4) Heavy metal accumulation by earthworms is particularly effective when
metals cannot be absorbed across cell membranes. Moreover, earthworms can benefit
plants with their vermicasts containing amylase, cellulase and chitinase, which
degrades organic matter and releases the nutrients for plant growth (Sinha et al., 2008).
Earthworms are suitable for heavy metal and radionuclide remediation in soil because
of (1) earthworms live in soil and contact with soil constantly, (2) earthworms can live
in many different types of soil and horizons, (3) earthworms can survive in
contaminated soil and as a result their body concentrations indicate bioavailability of
contaminants, (4) earthworms contact and uptake contaminants directly from their
skin, (5) earthworms consume soil as a food source therefore take up contaminants in
soils, (6) individual earthworms have enough biomass so contaminant concentrations
can be determined, (7) earthworm physiology and metabolism of metals are well
understood (Lanno et al., 2004), (8) earthworms can increase the mobility and
bioavailability of heavy metals to plants (Wen et al., 2004b).
Earthworms can take up heavy metals, pesticides, and organic contaminants such as
PAHs (poly aromatic hydrocarbons) (Contreras-Ramos et al., 2006, Sinha et al., 2008).
Using the bioaccumulation factor (BAF) it is possible to determine the bioavailability
of
the
contaminants
in
soils.
(BAF=concentration
of
contaminants
in
worm/concentration of contaminants in soil) (Fründ et al., 2011).
The bioaccumulation of earthworms is affected by biotic and abiotic factors. In the
soil, the accumulation depends on the concentration and physicochemical
30
bioavailability of contaminants. Neuhauser et al. stated that there was a significant
correlation between metal concentration (Cd, Zn, Pb, Cu but not Ni) in the soil
(sewage
sludge
applied)
and
the
concentration
of
these
in
earthworms
(Aporrectodeatuberculata and L. rubellus) (Neuhauser et al., 1995). The correlations
are Cd (R2=0.72), Cu (R2= 0.65), Cr (R2= 0.54), Pb (R2= 0.51), Zn (R2= 0.47), and Ni
(R2= 0.45) (Tischer, 2009). The chemical form also affects the accumulation. Lead as
a salt added to the soil is taken up more than aged contaminated Pb. The pH and redox
condition of soil strongly impact the speciation of contaminants (Alberti et al., 1996).
The microcompartment distribution (roots, organic matter, minerals, aggregates, and
soil water) can influence the accumulation (Ernst et al., 2008, Morgan and morgan,
1999).
Earthworms accumulate contaminants passively. The heavy metals or radionuclides
can enter earthworms through the body wall, mouth, and intestinal wall. The diffusion
is based on the difference between the concentration in pore water and earthworm
tissues (Jager et al., 2003). The factors in earthworms such as behavior (avoidance,
feeding, habitat preference, and spatial mobility) and physiology (cellular uptake,
physiological demand (regulation), binding proteins, granule formation, and excretion)
all contribute to the concentration of heavy metals in earthworms (Fründ et al., 2011).
Earthworms can affect the abiotic and biotic aspects of soil. The abiotic affect is
burrowing activity. Through burrowing, earthworms increase the aeration of soil by
creating channels within the soil, breaking down the soil into smaller particles, so that
microorganisms have more exposure space to degrade contaminants. As they burrow,
earthworms consume and digest organic matter containing contaminants into a smaller
sizes suitable for microbial degradation and remediation. The biotic effect is the
increase in microbial population (bacteria, fungi, and actinomycetes). This is because
the earthworm excretion contains urine, intestinal mucus, glucose, and nutrient (Sinha
et al., 2009)
Earthworms can change the bioavailability of pollutants as a result of stimulation of
the soil microbial population, alteration of soil pH and dissolved organic carbon
31
(DOC); and metal speciation and sequestration within earthworm tissue (Sizmur and
Hodson, 2009). The increased biomass of bacteria, actinomycetes, and fungi in
earthworm casts can lead to more metals being available to plants (Went et al. 2004).
This may be explained by the fact that microbial populations increase the degradation
of metal binding organic matter and release metals into soil solution (Rada et al.,
1996). In addition, earthworm activities improve aeration, organic matter content and
water availability (Tiunov and Scheu, 1999). Earthworms can change soil pH. Several
authors state that earthworms increase the availability of metals by decreasing the pH
of soil (El-Gharmali, 2002, Kizilkaya, 2004, Yu et al., 2005). However, many other
studies suggested that earthworms increase soil pH because of mucus excretion
(Schrader, 1994) while observing increasing in availability of metals (Ma et al., 2002,
Ma et al., 2003, Wen et al., 2006, Udovic and Lestan, 2007, Udovic et al., 2007).
Changes in soil DOC influence heavy metal availability. This is because DOC can
easily create complexes with heavy metals in soil solution and extract metals from the
soil surface with high affinity. Earthworms were found to produce humic acid
(Businelli et al., 1984). Humic acid was proven to enhance the availability of heavy
metals to plants by forming organo-metal compounds (Halim et al., 2003, Evangelou
et al., 2004). Moreover, the formation and release of metal chelating organic material
by earthworms can increase plant metal uptake (Currie et al., 2005, Wang et al., 2006,
Wen et al., 2006, Udovic et al., 2007).
Earthworms accumulate heavy metals in chloragogenous tissue around the posterior
alimentary canal (Morgan and Morris, 1982). There are two ways of maintaining
metals in this tissue. The first one (for Pb and Zn (Morgan and Morgan, 1989))
relating to the binding of insoluble metals, O-donating, and phosphate-rich granules
(chloragosomes) (Morgan and Morris, 1982, Morgan and Morgan, 1989, CotterHowells et al., 2005). The second one dealing with the binding of metals to low
molecular weight, S-donating ligands as metallothionein (Morgan and Morris, 1982,
Stürzenbaum et al., 1998, Cotter-Howells et al., 2005, Demuynck et al., 2007,
Demuynck et al., 2006). Cd-metallothioneins and Pb-metallothioneins were found in
L. terrestris and D. rubidus respectively by Ireland (1979) (Ireland, 1979). Earthworm
metallothionein isoform 1 wMT-1 is responsible for essential heavy metals such as Zn
32
and Cu at non toxic concentration. wMT-2 responses for non essential metals and
essential metals (Zn) at a higher toxic threshold level (Morgan and Morgan, 1989,
Sturzenbaum et al., 2001).
Eisenia andrei (Figure 1.5) was chosen
because it is easy to cultivate in laboratory
conditions
and
there
was
abundant
information of chemical toxicity on this
species (cited in (Lee et al., 2008)).
Lourenço et al, (2011) used Eisenia andrei
to assess the effects of heavy metals and Figure 1.5 Eisenia andrei
radionuclides
(uranium
mining)
on Source: http://www.thegardenforums.org
histopathology.
They
found
that
the
contaminated soil posed a risk to the fitness and survival of epigeic populations of E.
andrei. Effects on tissues paralleled the accumulation of heavy metals and
radionuclides (Lourenço et al., 2011). Eisenia andrei was also used by Natal-da-Luz et
al (2011) to investigate effects of Cr, Cu, Ni, and Zn contamination in sludge and
freshly spiked soils. They concluded that the decrease in bioavailability of the metal
was promoted by the increase of organic matter (Natal-da-Luz et al., 2011).
Furthermore, E. andrei was also used by Saxe et al. (2001) to create a model
describing the relationship between Cd, Cu, Pb and Zn body concentration in the
worms with soil pH, metals, and organic carbon in order to predict concentration of
metals in contaminated sites using earthworms as biomonitors (Saxe et al., 2001).
1.7 Use of toxicity tests for environmental risk assessment of contaminated soils
Ideally for toxicity testing all ecological and commercial organisms that are associated
with the soil should be used. However, this is impossible and certain tests are selected
to estimate the ecosystem risk as much as possible. The test organisms must be
representative for selected environment and sensitive to contaminants. Toxicity testing
with several test species has been standardized by several organization including
OECD, ISO, ASTM and EPA. The toxicity data can be accessed through biochemical,
33
physiological, reproductive, behavioral effects, lethality, reproduction, and growth
(Stephenson et al., 2002).
Earthworm toxicity tests are the most popular. Two major tests based on mortality and
production of Eisenia sp. have been standardized by OCED (OECD, 2004a).
Avoidance, weight loss, and chemical bioaccumulation can also be used to estimate
the toxicity of chemicals in the terrestrial environment (Domínguez, 2008).
Plants are important and they reflect the quality of soils. Several methods have been
evaluated to estimate the toxicity of contaminants to plants. The endpoints are survival
or growth (change in height/ length or biomass), germination and growth of seeds,
specific enzymes, respiration (total and dark) (OECD, 2006a).
Microbial toxicity tests are also available to assess the quality of the soil.
Microorganisms have important functions in terrestrial ecosystems such as organic
matter decomposition and nutrient cycling. The targets of these tests are soil
respiration, nitrogen mineralization, and soil enzymes which are dehydrogenases, βglucosidases, ureases, phosphatases, arylsulphatases, cellulases, and phenol oxidases
(Dick et al., 1996).
To assess the toxicity of soil leachates, the bioluminescence assay with Vibrio fischeri)
(Microtox) is used. This marine bacterium is luminescent and sensitive to a variety of
heavy metal compounds (Johnson et al., 1942). In this test the effects of contaminants
are reflected by the light emission which is measured by a photometer in an interval of
time. Due to the fact that this bacterium has fast metabolism, this assay gives a rapid
result on the toxicity of leachates. This test is widely used and considered as sensitive,
reproductive and precise (Domínguez, 2008).
34
1.8 Integrated soil microcosms
Single species toxicity tests which
have been commonly used provide
inadequate data to predict the total
effect of the toxicants under the field
conditions
(Edwards,
2002).
An
integrated soil microcosm (ISM) in
which multiple species are included
can provide much more information
about the interactive impact of
toxicants on biological activity. It
can mimic and help to estimate Figure 1.6 Main components of an integrated
correctly the impact of the hazards soil microcosm (ISM)
under field conditions. Moreover, a variety of microcosms have been used for
ecototoxicological assays due to the advantages of being easy to operate, being
inexpensive, and controllable under laboratory conditions with several replicates
(Chen and Edwards, 2001).
The integrated soil microcosm is constructed as a cylinder, generally of plastic,
polyvinyl chloride (PVC) or high-density polyethylene (HDPE). Inside it contains
field soil which is sieved and thoroughly mixed with indigenous soil organisms
(Edwards, 2002) (Figure 1.6). Approximately 1 kg of air dried and homogenized soil is
filled in each ISM. Together with indigenous organism earthworms (1.5 g) and plants
are introduced into an ISM. At the bottom of each ISM ion exchange resins are placed
separately with the bottom of soil core by a thin glass wool layer. This is used for
collecting leachates and preventing plant roots from going out. Irrigation is conducted
two or three time a week at a level of 40-60% of field water holding capacity.
Leachates collection is implemented once a week by adding excess water. The funnel
will be installed to replace the ion exchange layer if the study relating to the fate and
leaching of a pesticide. Leachate from microcosms will be collected in a plastic
container for further analysis (Edwards, 2002). Though ISMs have been used for
35
testing the toxicity of environmental pollutants there are few published studies on the
use of ISMs to evaluate bioremediation of trace metal pollutants.
1.9 Aims of the project
This project aimed to evaluate the efficacy of bioremediating Cs and Pb in
contaminated soils using the selected biota. In order to accomplish the aim the
following objectives were addressed:
1. The accumulation of Cs from soil contaminated with Cs only by sunflowers
with and without earthworms
2. The accumulation of Pb from contaminated soil by sunflowers with and
without earthworms.
3. The accumulation of Cs and Pb from soil contaminated with a mixture of
these metals by sunflowers with and without earthworms.
4. The effects of Cs and Pb as individual metal contaminants and a mixture of
Pb and Cs on plants and earthworms in contaminated soil.
5. The fate of Cs and Pb in the soil profile through their mobility, leaching and
effects on soil pH, conductivity and organic matter.
The innovationin this study is the use of Eisenia andrei species and the sunflower
Helianthus annuus (dwarf-sensation) in combination; The application of Cs and Pb as
a mixture; and the use of the integrated soil microcosms.
36
Chapter 2. Materials and methods
Experiments
were
conducted
in
controlledenvironmental
conditions
in
the
Ecotoxicology laboratory in Bundoora West Campus. Integrated Soil Microcosms
were used for all experiments and natural uncontaminated soil was used as a medium.
A population of earthworms (Eisenia andrei) and the plant dwarf-sunflower
(Helianthus annuus) were used as test organisms (hereafter referred to as “biota”). The
trace metals Cs and Pb were used to spike the natural uncontaminated soil. Both
metals were tested individually and also tested as a mixture. Two different, complete
and independent experiments were conducted each running for 4 weeks.
2.1. Soil collection
Natural uncontaminated soil was collected from a construction site near the
Ecotoxicology laboratory, Bundoora West Campus, RMIT University. Soil after
collection was brought to the Ecotoxicology laboratory in plastic containers and kept
in a cool room at 4oC until use. The soil was then air-dried, gently crushed, and sieved
through a 5-mm stainless-steel mesh to remove unwanted plants and coarse stones.
Then, the physical characteristics of soil namely moisture, pH, conductivity, organic
matter, organic carbon, water holding capacity and soil texture were measured. Cs,
Pband potassium content of soil was also measured as detailed below.
2.2 Soil physical and chemical characterization
Physical and chemical characteristics of soil were measured before the start of
experiments and also at each sampling time. Soil moisture was measured following
ISO 11465 (1993) based on the measurement of weight loss of soil at 105oC(ISO
11465, 1993). First of all, 5 – 10 g soil was transferred to a dried and tarred crucible of
known weight. Soil was then dried at 105°C for 24 hours, cooled in desiccators and
weighed. The moisture content in weight percentage was calculated as follow:
Moisture (%) = ((A – B)/ B – crucible weight)*100%.
Where A : Weight of crucible and soil sample (gram).
B : Weight of crucible and oven-dried soil sample (gram).
37
The pH and conductivity of the soil were measured in a soil - water suspension (1: 5,
w/v extraction ratio) according to the method described in FAO (1984). Firstly, 5 gram
of the soil sample was mechanically shaken in polypropylene flasks with 50 ml of
deionized water for 30 minutes. The mixture was then allowed to rest for 1 h and the
pH of the overlying solution was measured using a TPS WH-81 pre-calibrated pHconductivity-salinity meter. Conductivity was measured in the same suspension after
leaving it to rest overnight, to allow the bulk soil to settle. (FAO, 1984). The loss-onignition (LOI) method was used to determine organic matter content. A known weight
of soil sample was placed in a ceramic crucible which was then heated to 440°C
overnight (Blume et al., 1990, Nelson and Sommers, 1982). The samples were then
cooled in desiccators and weighed. Organic matter content (OMC) was calculated as a
percentage of the difference between the initial and the final sample weight divided by
the initial sample weight. All weights were corrected for moisture/water content prior
to organic matter content calculation. To measure water holding capacity soil samples
were placed in polypropylene flasks and immersed in water for 3 h. After this period
samples were drained for 2 h by rejecting exceed water with absorbent paper. The
Water Holding Capacity (WHC) was determined by weighing each replicate before
and after drying at 105o C, until weight stabilization (ISO, 2008).
The textural classification of a soil sample was determined by measuring the relative
amounts of sand, silt, and clay particles, then using a soil triangle to determine the soil
type. The comparative volumes of sand, silt, and clay were determined based upon the
fact that the different sized particles will settle out of a mixture at different rates.
Initially, a 40 to 50 mL of soil sample was placed in the graduated cylinder. Then
water was added water until the total volume of soil and water is about 80-100 mL.
The top of the graduated cylinder was covered with a piece of plastic wrap and secured
with a rubber band. The cylinder was inverted several times until the soil was
thoroughly suspended in the water. The cylinder was shaken to mix the water and soil
thoroughly. The cylinder was then placed on the table and the soil material allowed to
settle for at least 30 minutes. The different soil materials settled to the bottom at
different rates depending upon their particle sizes: sand size > silt size > clay size. The
38
volume of the sand, silt and clay layers were estimated and recorded using the marks
on the graduated cylinder. The method states that there should be at least three
reasonably distinct layers in the graduated cylinder representing sand (bottom), silt
(middle) and clay (top). There might also be a dark humus layer above the clay layer,
or possibly floating on top of the water. The volumes of the three layers and the total
of volume of samples were recorded and the percent composition by volume for each
layer was calculated.
The concentration of Cs, K and Pb in soil were measured followed Kovacs et al (2000)
(Kovacs et al., 2000). Soil was oven dried at 60-80oC for 48 hours prior to weighing.
One gram of dried soil was weighed into a test tube. Then, 5 mL HNO 3 (65-70%) was
added. Soil was pre-digested on a heating block for 30 minutes at 60oC. Five milliliters
H2O2 was added slowly (1-2 ml at a time). Soil was further heated at 120oC for 4 hours
and 30 minutes and then the tubes removed from the heating block and cooled to room
temperature. The digested solution was filtered through Whatman No.52 filter papers
into 50 ml volumetric flasks. Test tube and filter paper were rinsed twice with
deionised water. The volumetric was then filled to the mark with deionised water and
stored at 4oC in suitable polyethylene tubes for analysis by either AAS or ICP-MS.
2.3 Integrated Soil Microcosms
To perform this research study, two experiments using an Integrated Soil Microcosm
(ISM) were conducted in order to assess the behavior, accumulation and mobility of
Cs and Pb in the ecosystem using plants, earthworms, soils and leachates as test
elements. Soil microcosms were set up in a laboratory with controlled environmental
conditions such as: air temperature 20ºC ± 2ºC, with a 16:8 light: dark cycle.The soil
microcosms used consisted of a polystyrene cup, with a diameter of 145 mm at the
base and height 140 mm filled with 1.7 kg of natural uncontaminated soil forming a 12
cm deep layer (total depth of soil layer: 12 cm). Such cups are impermeable,
lightweight, rigid and highly resistant to acids, bases, and biological degradation. A
bottom plate was used to collect leachate during the tests. Then, all soil cores were
placed in a metallic structure (Figure 2.1).
39
Figure 2.1 Integrated soil microcosms
The earthworms Eisenia andrei used to perform the tests were obtained from a
commercial warehouse. In the laboratory the earthworms were kept in a culture at
20+2oC with a 16:8 (light: dark) photoperiod and they were maintained on a diet of
horse manure. Only adult animals with a developed clitellum were selected for the
tests (50-60 mm length) (Figure 2.2).
Figure 2.2 Earthworm Eisenia andrei
The sunflower Helianthus annuus (dwarf-sunflower) was selected for seed
germination, growth assays and bioaccumulation studies. This plant was selected
because it has a fast growth and high biomass production and has been reported as
very sensitive and tolerant to a great number of contaminants, including metals and
radionuclides. Seeds were purchased from a local supplier and damaged seeds were
discarded after visual inspection. Germination and growth tests were performed
following standard procedures described in ISO 11269-2 guidelines(ISO, 1995)
(Figure 2.3).
40
Figure 2.3 Sunflowers (Helianthus annuus)
2.4 Experimental design
Both experiments were performed in the same laboratory under the same controlled
environmental conditions and the procedures and the set up was the same. Soil
microcosms were randomly assigned in the laboratory accordingly to the treatments
used including the control. An individual batch of natural uncontaminated soil was
prepared for each experiment. Then distilled water was added to each batch of soil to
establish a percentage between 20 and 40 of the maximum water holding capacity
(WHC). Soil was then left for two weeks for acclimatization and water loss due to
evaporation was monitored every second day and, when required it was replenished
with deionised water. After two weeks, 1.7 kg of wet soil was placed inside each soil
microcosms. To maintain the soil moisture content at the desired constant levels,
losses in soil microcosm weight due to evapotranspiration was compensated for on a
daily basis using a fixed volume of distilled water. Using a syringe, the solutions
contaning the metals were applied to the soil surface in the form of very small drops
accordingl to the experimental design and kept at 20±2oCfor two weeks in moist
condition before starting the experiments. Microcoms were colonized with plants and
earthworms according to the experimental designs (see Table 2.2 and 2.3) and kept for
4 weeks.
41
Ten earthworms (50-60 mm length) with a developed clitellum were introduced into
earthworms only and earthworms+plants microcosms. The weight of the earthworms
was recorded to calculate the gain/ loss biomass. Ten dwarf sunflower seeds were
sown in selected microcosms according to the experimental design. After germination,
the plants were stripped to three. The experiment started after the germination of 50%
of the seeds in the control. Since all effects within an ISM study are defined as
deviations from the untreated control, the latter is of utmost importance. Therefore,
there were four replicates for the control and also four replicates per sampling time for
the other treatment levels. At each sampling time, soil from each ISM was taken at
different soil depths, 0-5 cm and 5-10 cm, and all the earthworms found were collected
and counted. The following parameters were also measured at each sampling time
(Table 2.1)
Table 2.1 Parameters measured at each sampling time
Biota
Parameters
Compartment and size of sample
Survival
Whole soil core
Soil layer: 0-5 cm
Vertical distribution
Soil layer: 5-10 cm
Earthworms
Biomass
Whole soil core
Soil layer: 0-5 cm
Cs, K, and Pb concentration
Plants
Soil layer: 5-10 cm
Shoots and roots length (cm)
Cultivated plants on ISMs
Biomass (g)
Cultivated plants on ISMs
Cs, K, and Pb concentration in roots
Cultivated plants on ISMs
and shoots
Moisture,
Soils
organic
conductivity.
Cs,
matter,
K
concentration.
Leachates
pH, toxicity. Cs, K and Pb
concentration.
and
pH,
Pb
Soil layer: 0-5 cm
Soil layer: 5-10 cm
Collected at each sampling time.
At each sampling time leachate was collected from the bottom of each soil
microcosms. The three specimens of dwarf-sunflower were harvested, measured and
weighed. Shoots were cut within 1 cm of the soil level and soil brushed off prior to
weighing. Earthworms were sampled at the end of the experiments in both soil layers
42
by hand-sorting, weighed immediately after collection and placed on wet paper towels
overnight to clear their guts and then re-weighed.
2.4.1 Experiment 1: Caesium
In this experiment, soil microcosms were randomly assigned to three treatments
including the control. Soil microcosms were spiked with two concentrations of Cs
chloride (CsCl) (25 and 250 mg/kg). There were two sampling times at 2 and 4 weeks.
A total of 96 soil microcosms were used (Refer Table 2.2 for more details)
Table 2.2 Experimental design used for Cs
ISM
Sampling 1
Sampling 2
Number
(2 weeks)
(4 weeks)
No biota
1-8
1-4
5-8
Earthworms only
9-16
9-12
13-16
Plants only
17-24
17-20
21-24
Earthworms and plants
25-32
25-28
29-32
No biota
33-40
33-36
37-40
Earthworms only
41-48
41-44
45-48
Plants only
49-56
49-52
53-56
Earthworms and plants
57-64
57-60
61-64
No biota
65-72
65-68
69-72
Earthworms only
73-80
73-76
77-80
Plants only
81-88
81-84
85-88
Earthworms and plants
89-96
89-92
93-96
Treatments
Control
100%
natural soil
25 mg/kg Cs
250 mg/kg Cs
Total
96
2.4.2 Experiment 2: Cs and Pb as a mixture of contaminants.
In this experiment, three treatments including the control were prepared. Some soil
microcosms were spiked with 1500 mg/kg Pb (PbNO3) solution and other microcosms
were spiked with a mixture of Cs and Pb having 1500mg/kg Pb and 250mg/kg Cs.
There was only one sampling time at 4 weeks, based on the results of experiment 1
43
where there was no significant difference in the results between 2 and 4 weeks. A total
of 48 soil microcosms were used (refer Table 2.3 for more details)
Table 2.3 Cs and Pb accumulation experimental design
ISM
Treatments
Control
100% natural clean
soils
1500 mg/kg [Pb]
1500 mg/kg [Pb] +
250 mg/kg [Cs]
Number
No biota
1-4
Earthworms only
5-8
Plants only
9-12
Earthworms and plants
13-16
No biota
17-20
Earthworms only
21-24
Plants only
25-28
Earthworms and plants
29-32
No biota
33-36
Earthworms only
37-40
Plants only
41-44
Earthworms and plants
45-48
Total
48
2.5 Soil sampling, digestion and chemical analysis
Each soil core was cut into 2 layers. In each soil layer the following parameters: soil
moisture, pH, conductivity, and organic matter were measured as described above.
Soil digestion and chemical analysis was conducted as described in section 2.2.
2.6 Leachate assessment
Leachates were collected at each sampling time. They were collected to measure pH,
toxicity and heavy metal concentration. This indicated the potential of groundwater
contamination. To estimate heavy metal concentration in leachates, 2% (v/v) of
concentrated HNO3 was added into leachates which were then analyzed for heavy
metals using atomic absorption spectrometry (Olowu et al., 2009).
2.7 Earthworm sampling, digestion and chemical analysis
Earthworms were collected and counted in both soil layers of the soil core to measure
the behavior and survival of earthworms. Then, they were weighed to obtain
44
earthworm biomass and analysed for heavy metal concentration. Prior to acid
digestion, earthworms from each layer were gently washed with distilled water to
remove surface adherent soil particles and placed in petri dishes with moist filter paper
(Whatman No.1). Petri dishes were kept in the dark at room temperature (19-22oC) for
three days to allow the earthworms to purge their digestive tract. The filter papers were
changed at the end of the first and second day to reduce coproghagy and/ or fungal
growth. On the third day the worms were depurated for a further day in a moist
environment in petri dishes with a few drops of distilled water with no filter paper, to
permit the egestion of filter paper from the gut. After the defecation period, the worms
were individually placed in pre-weighed acid washed glass test tubes and killed by
freezing. Next the test tubes containing the soil-purged worms were oven dried at 80oC
for a minimum of 36 hours until they had reached constant weight. The test tubes were
then placed in a dry block heater at 100oC for 2 hours. A few drops of concentrated
HNO3 were added. After 15 minutes, 2ml of concentrated HNO3 was added into each
test tube. H2O2 (0.5 mL) was also added to assist digestion. After acid digestion had
taken place, the samples were left to cool to room temperature (19-22oC). The samples
were then filtered through Whatman No.1filter paper into a 50 mL volumetric flask.
All samples were made up to 50 mL with distilled water. Procedural blanks were
carried throughout the digestion and analysis to maintain quality control(Rodrigues,
1997). Finally atomic absorption spectrometry (Varian 220) and ICP-MS was used to
determine Pb, Cs and potassium content.
Transfer factor (TF) and total uptake (M) of Cs and Pb were calculated similar to that
for plants (see section 2.8)
2.8 Plant sampling, wet digestion and chemical analysis
At the end of the assays, plants were harvested from all replicates. Before analysis,
roots and shoots of each plant were carefully washed with tap and deionised water to
remove any soil and surface dust. Then roots and shoots of plants were separated to
measure the length and biomass. Plant material was then dried at 70oC, weighed,
ground and digested using a mixture of acids (Tüzen, 2003). One gram of <2 mm
fraction of plant samples were weighed into test tubes. Then 3 mL concentrated HNO3
and 1 mL concentrated HCl were added. Samples were heated at 85 oC for 3 hours in
45
the heating block and cooled to room temperature. They were then filtered through
Whatman filter paper No. 1 into 25 ml volumetric flasks and both test tubes and filter
papers were washed into the volumetric flask. Digested solutions were made up to 25
mL with deionized water and then stored in polyethylene tubes at 4oC until analysis by
Atomic absorption Spectrometry (Varian 220). Background correction was used in Pb
analysis but not in Cs analysis as recommended by the manufacturer’s protocol.
Pb concentration in plants together with Pb concentration in soils was used to calculate
Bioaccumulation Coefficients (BACs) for plant(Li et al., 2007, Cui et al., 2007). BACs
are the relationship between the concentrations of Pb in the plant and the concentration
of the same element in the soil (Equation 1). For Cs, transfer factor with the same
definition was used instead of Bioaccumulation Coefficients by Cook et al. (Cook et
al., 2007). In order for ease of comparison, the transfer factor was used to address the
relationship between Cs and Pb in soil and plant shoots. Translocation Factor (TLF)
was described as ratio of Cs and Pb in plant shoot to that in plant root given in
equation 2 (Cui et al., 2007; Li et al., 2007). Cs and Pb total uptake (extracted) M
expressed the amount of Pb was extracted from soils (equation 3) (Wuana and
Okieimen, 2011).
TF= [Metals]shoot / [Metals]soil
(1)
TLF = [Metals]shoot/ [Metals]root (2)
M
=[Metals]plant tissues x Biomass
(3)
2.9 Data analysis
Data was collected to compare the following parameters statistically:
The difference in the length of shoots and roots and difference in biomass of plants
between treatments. The differencein biomass, survival and movement of earthworms
between treatments. The difference in pH of leachates between treatments. The
difference between treatments in the Cs and Pb concentrations in plants, earthworms,
soil and leachates. Statistical analysis of the data was performed using the SPSS 14.0
statistical software. The data were tested for normality and equality of variances. The t
test, one way and two way Analysis of Variance (ANOVA) were performed, followed
by the Tukey’s test for multiple comparisons as appropriate when data were normally
distributed with equal variances;the Kruskal-Wallis test followed by the Mann46
Whitney test were performed on data that did not fulfill these criteria and means
considered significantly different from each other at P < 0.05. All data was expressed
as mean ± standard error.
47
Chapter 3. Results
3.1 Physical and chemical characteristics of soil
Table 3.1describes the general physical and chemical characteristics of the natural
uncontaminated soil used to perform the experiments. According to the textural
characterization the soil was classified as a clay loam soil.
Table 3.1 Properties of clean soil collected near the Ecotoxicology laboratory, RMIT
Bundoora west campus. (mean ± SE, n=3)
Parameters
Clean soil
7.2±0.0 for Cs experiment
pH
6.60 ± 0.03 for Cs+Pb experiment
Conductivity (μS/cm)
77.40±0.05
Moisture (%)
25.00±0.28
Organic matter (%)
11.00±0.03
Organic carbon (%)
18.96±3.20
Water holding capacity (%)
Particle size
(clay loam)
Potassium concentration (mg/kg)
31.0±0.5
Sand
21.6%
Silt
42.1%
Clay
36.3%
2,389
Cs concentration (mg/kg)
1.80±0.08
Pb concentration (mg/kg)
15.00±0.28
3.2 Caesium experiment
3.2.1. Soil parameters after 2 and 4 weeks in experimental ISMs
Results revealed that all pH values were alkaline after two and four weeks in the
experiments regardless of treatments, the presence of biota, or whether the soil was
from the top or bottom layer (Table 3.2). Soil conductivity had changed at the end of
the experimental period. However, these values did not exceed the range from 0 to
2000 μS/cm which was classified as within the conductivity range which causes no
adverse effects on plants (Table 3.3). Soil moisture was recorded to be between 2640% in top layer soils and 32-43% in bottom layer soils. The moisture of the bottom
48
layer soils was higher than that of the top layer soils (Table 3.4). Organic matter
content of soil remained between 11-12% across treatments, with and without biota
and in soil layers (Table 3.5).
Table 3.2 Soil pH after 2 and 4 weeks in experimental ISMs( mean ± SE, n=3)
Treatment
Sampling
times
Biota
Soil
layers
No biota
Earthworms
Plants only
Worms+Plants
only
2 weeks
Control
4 weeks
2 weeks
25 mg/kg
4 weeks
2 weeks
250 mg/kg
4 weeks
Top
7.5±0.0
7.5±0.0
7.7±0.1
7.6±0.0
Bottom
7.5±0.1
7.5±0.0
7.6±0.0
7.6±0.0
Top
7.0±0.1
7.3±0.0
7.4±0.0
7.3±0.1
Bottom
7.2±0.1
7.4±0.1
7.4±0.0
7.3±0.1
Top
7.7±0.0
7.7±0.0
7.7±0.0
7.7±0.1
Bottom
7.7±0.1
7.5±0.0
7.7±0.1
7.6±0.0
Top
7.4±0.0
7.3±0.1
7.4±0.0
7.4±0.0
Bottom
7.5±0.0
7.4±0.1
7.5±0.1
7.4±0.1
Top
7.3±0.0
7.4±0.1
7.3±0.0
7.2±0.1
Bottom
7.3±0.0
7.3±0.0
7.4±0.0
7.3±0.1
Top
7.4±0.1
7.4±0.0
7.4±0.0
7.4±0.1
Bottom
7.5±0.2
7.5±0.1
7.5±0.0
7.5±0.1
Table 3.3 Soil conductivity (μS/cm) after 2 and 4 weeks in experimental ISMs
(mean ± SE, n=3)
Treatment
Sampling
Soil
times
layers
Biota
No biota
Earthworms
only
2 weeks
Control
4 weeks
2 weeks
25 mg/kg
4 weeks
2 weeks
250 mg/kg
4 weeks
Plants
Worms+Plants
only
Top
49.1±1.9
62.2±4.7
46.6±3.8
63.1±13.5
Bottom
60.8±8.8
68.8±3.6
61.7±5.9
67.0±10.1
Top
57.5±3.0
68.1±0.3
45.9±1.5
73.7±29.6
Bottom
60.8±0.6
61.9±21.7
63.4±1.9
74.8±10.8
Top
53.7±2.5
73.2±1.1
61.1±7.8
60.0±1.7
Bottom
70.1±9.8
82.1±1.8
69.0±0.4
80.8±7.4
Top
55.7±4.7
61.1±4.2
61.3±1.3
65.4±8.1
Bottom
56.1±8.8
81.4±10.0
68.7±7.3
72.9±5.6
Top
71.0±5.0
87.9±7.1
83.6±8.5
87.5±15.1
Bottom
86.3±19.1
99.0±4.8
78.2±5.2
86.7±10.8
Top
76.9±21.1
85.9±8.1
75.5±10.4
86.8±6.6
Bottom
69.6±1.1
84.5±13.4
70.7±7.8
92.1±13.5
49
Table 3.4 Soil moisture (%)after 2 and 4 weeks in experimental ISMs (mean ± SE,
n=3)
Treatment
Sampling
Soil
times
layers
2 weeks
Control
4 weeks
2 weeks
25 mg/kg
4 weeks
2 weeks
250 mg/kg
4 weeks
Biota
No biota
Earthworms
Plants
Worms+Plants
only
Top
35.7±6.3
29.1±0.2
29.3±0.1
29.6±0.1
Bottom
37.3±5.1
34.1±0.6
33.8±2.9
35.0±2.4
Top
25.7±0.3
27.1±0.2
27.1±1.0
27.2±1.1
Bottom
35.9±0.9
34.3±0.2
39.6±0.9
35.3±0.9
Top
32.2±1.2
31.1±0.8
31.7±1.1
33.2±0.5
Bottom
38.3±2.8
37.2±0.0
39.1±1.1
38.7±0.9
Top
27.3±1.7
30.1±0.0
30.3±2.8
30.3±0.7
Bottom
41.2±2.8
38.6±0.1
37.0±0.3
38.9±0.4
Top
35.2±0.6
37.5±0.3
33.6±2.0
36.2±0.5
Bottom
41.8±1.4
41.6±0.6
40.3±3.2
40.7±2.1
Top
33.7±0.2
34.3±1.0
32.3±0.3
33.5±0.3
Bottom
38.3±0.9
38.3±0.9
35.4±1.7
36.8±1.0
Table 3.5Soil Organic Matter Content (%) after 2 and 4 weeks in experimental ISMs
(mean ± SE, n=3)
Treatment
Sampling
Soil
times
layers
Biota
No biota
Earthworms
Plants
Worms+Plants
only
2 weeks
Control
4 weeks
2 weeks
25 mg/kg
4 weeks
2 weeks
250 mg/kg
4 weeks
Top
12.5±0.2
12.1±0.0
12.4±0.5
12.7±0.3
Bottom
12.5±0.1
12.6±0.2
12.4±0.3
12.3±0.1
Top
11.4±0.1
11.6±0.3
11.6±0.1
11.3±0.1
Bottom
11.5±0.1
11.7±0.3
11.7±0.2
11.7±0.5
Top
12.7±0.0
12.5±0.4
12.2±0.3
11.9±0.1
Bottom
12.3±0.5
12.7±0.1
12.6±0.5
12.0±0.3
Top
11.2±0.4
11.2±0.2
11.6±0.2
11.0±0.2
Bottom
11.4±0.3
11.1±0.0
11.3±0.4
11.3±0.1
Top
9.8±0.0
10.3±0.1
10.5±0.2
10.4±0.1
Bottom
9.6±0.0
10.6±0.1
10.8±0.0
10.6±0.0
Top
11.7±0.1
11.7±0.3
12.0±0.0
11.9±0.1
Bottom
11.7±0.1
12.0±0.0
11.9±0.3
11.9±0.2
50
3.2.2 Caesium in soils after 2 and 4 weeks
The results show that Cs was higher in soils in both 250 mg/kg and 25 mg/kg
treatments than in controls regardless of whether they were soils from the top or
bottom layer (Figure 3.1). The non-parametric Kruskal-Wallis test was used.
In the 25 mg/kg treatment:
At top layer soils, after two weeks, there was a significant difference in Cs in soil with
different biota (P=0.048). The concentration of Cs in soils without biota (33.1 ± 2.0
mg/kg) were higher than in soils with earthworms and plants (41.3 ± 0.9 mg/kg)
(P=0.021). After 4 weeks, there was no significant differences in soil with different
biota (P=0.203). There was a significant reduction of Cs in soils from 2 weeks to 4
weeks sampling in the 25 mg/kg treatment (P<0.001). The Cs concentrations were
36.4 ± 1.2 mg/kg and 29.1 ± 0.7 mg/kg respectively.
In bottom layer soils, after 2 weeks, there was a significant difference in Cs
concentration between soil with different biota (P=0.03) which were between the soils
without biota and with earthworms only (P=0.02) and between soils with earthworms
only and soils with plants alone (P=0.043). The Cs concentration of soils with
earthworms only (5.4 ± 0.3 mg/kg) was higher than in soils without biota (4.5 ± 0.1
mg/kg) and soils with plants only (4.6 ± 0.3 mg/kg).
After 4 weeks, there was also a significant difference between soils with different biota
(P=0.004). The Cs concentrations in soils were earthworms only > earthworms+plants
> plants only > no biota (4.9 ± 0.2 mg/kg, 4.1 ± 0.1 mg/kg, 2.8 ± 0.4 mg/kg, 2.1 ± 0.2
mg/kg). There was also a significant difference in Cs concentration in soils from 2 to 4
weeks (P<0.001). Cs concentration was reduced from 4.9 ± 0.1 mg/kg to 3.5 ± 0.3
mg/kg after 4 weeks (Figure 3.1).
51
In the 250 mg/kg treatment:
At top layer soils, at both 2 and 4 weeks sampling times, there were no significant
difference in Cs concentration between soils with different biota (P=0.553 for 2 weeks
and P=0.182 for 4 weeks). There was also no significant difference between the first
and the second sampling (P=0.066).
In bottom layer soils, after 2 weeks, there was a significant difference between soils
with different biota namely between soils with no biota when compared with soils with
only earthworms (P=0.043) and soils with earthworms+plants (P=0.043); between
soils with plants alone when compared with soils with only earthworms (P=0.021) and
soils with earthworms+plants (P=0.021). The Cs concentration of soils with only
earthworms (37.4 ± 5.0 mg/kg) or soils with earthworms+plants (49.2 ± 6.9 mg/kg)
were higher than in soils without biota (13.1 ± 7.5 mg/kg) or plants only (14.7 ± 3.3
mg/kg). After 4 weeks, there was no significant difference in soil Cs with different
biota (P=0.402).
There was no significant difference in Cs in soils from 2 to 4 weeks (P=0.164) (Figure
3.1).
52
2 weeks
250W+P
250P
250W
250N
25W+P
25P
25W
25N
CW+P
CP
CW
CN
250W+P
250P
250W
250N
25W+P
25P
25W
Bottom layer
25N
CP
CW
Top layer
CW+P
450.0
400.0
350.0
300.0
250.0
200.0
150.0
100.0
50.0
0.0
CN
Cesium concentration (mg/kg)
Cesium concentration in soils
4 weeks
Treatments
Figure 3.1 Cs concentrations in soils after 2 and 4 weeks
CN: Control-No biota; CW: Control-Earthworms only; CP: Control-Plants only; CW+P: Control-Earthworms+Plants. 25N: 25mg/kg Cs-No biota; 25W: 25mg/kg Cs-Earthworms
only; 25P: 25mg/kg Cs-Plants only; 25W+P: 25mg/kg Cs-Earthworms+Plants. 250N: 250mg/kg Cs-No biota; 250W: 250 mg/kg Cs-Earthworms; 250P: 250 mg/kg Cs-Plants;
250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
52
3.2.3 Effects of Cs on leachates
3.2.3.1 Leachate pH
Table 3.6 pH of leachates after 2 and 4 weeks (mean ± SE, n=4)
Treatment
Control
25 mg/kg
250 mg/kg
Biota
Sampling
times
No biota
Worms
Plants
Worms+Plants
2 weeks
7.8±0.04
7.6±0.04
7.7±0.08
7.5±0.08
4 weeks
7.6±0.04
7.71±0.07 8.03±0.11
7.42±0.09
2 weeks
7.95±0.06
7.55±0.1
7.63±0.11
7.6±0.07
4 weeks
7.86±0.08
7.5±0.05
7.64±0.05
7.28±0.06
2 weeks
7.46±0.02 6.94±0.09
7.5±0.04
7.02±0.2
4 weeks
6.9±0.14
7.29±0.08 7.18±0.02 6.99±0.04
After 2 weeks, there was a significant difference between treatments (P<0.001) namely
between the 250 mg/kg treatment and the control (P<0.001) and 25 mg/kg (P<0.001).
Leachate pH in the 250 mg/kg treatment was lower than in the control and in the 25
mg/kg treatment. There was a significant difference between leachate with different
biota across treatments (P=0.019) which were between the leachates in the treatment
without biota in comparison with the leachates from the treatment with only
earthworms (P=0.013) and with earthworms+plants (P=0.01). The pH of leachates in
the treatment without biota was higher in both cases. Within treatments, there was a
significant difference in the leachates from the control (P=0.049) and 250 mg/kg
(P=0.028) treatment. In the control, pH of the leachates without biota were higher than
in the leachates from the treatment with only earthworms (P=0.027) and
earthworms+plants (P=0.027). In 250 mg/kg treatment, pH of the leachates from the
treatment without biota and plants only were higher than the pH of the leachates from
the treatment with only earthworms (P=0.019 and P=0.020) (Table 3.6).
After 4 weeks, there was a significant difference in leachates between treatments
(P<<0.001) which were between 250 mg/kg compared with control (P<0.001) and 25
mg/kg (P<0.001). The pH of leachates collected from 250 mg/kg was lower than in the
leachates from the control and 25 mg/kg. There was no significant difference between
53
leachates with different biota across treatments (P=0.0.54). Within treatments, there
was a significant difference in the control (P=0.011), 25 mg/kg (P=0.006), and 250
mg/kg (P=0.009) (Table 3.6).
In the control, pH of leachates with only plants were higher than in leachates in the
treatment without biota (P=0.019), only earthworms (P=0.038) and earthworms+plants
(P=0.020). In the 25 mg/kg treatment, the leachates from the treatment without biota
had a higher pH than the leachates from the treatment with only earthworms
(P=0.020), and earthworms+plants (P=0.020). The leachates from the treatment with
plants alone has a higher pH than those from treatments with earthworms+plants
(P=0.019). In the 250 mg/kg treatment, leachates from the treatment without biota had
a lower pH than the leachates from the treatment with earthworms alone (P=0.019),
plants alone (P=0.034). Leachates from the treatment with earthworms alone had a
higher pH than those from the earthworms+plants ISMs (P=0.019) (Table 3.6).
From 2 to 4 weeks, there was no significant change of pH in all treatments (P=0.446,
P=0.086, and P=0.236 for control, 25 mg/kg and 250 mg/kg respectively) (Table 3.6).
3.2.3.2 Caesium concentration in leachates
Caesium was only detected in the leachates from the 250 mg/kg treatment. There was
no significant difference between the first and second sampling (P=0.489). In each
sampling time, there was no significant difference in Cs concentration between
leachates with different biota (P=0.627) (Figure 3.2).
54
2 weeks
250W+P
250P
250W
250N
25W+P
25P
25W
25N
CP
CW+P
CW
CN
250W+P
250P
250W
250N
25W+P
25P
25W
25N
CW+P
CP
CN
0.4
0.35
0.3
0.25
0.2
0.15
0.1
0.05
0
CW
Cesium concentration (mg/L)
Cesium in Leachates
4 weeks
Treatments
Figure 3.2 Caesium concentrations in leachates after 2 and 4 weeks
CN: Control-No biota; CW: Control-Earthworms only; CP: Control-Plants only; CW+P: ControlEarthworms+Plants. 25N: 25mg/kg Cs-No biota; 25W: 25mg/kg Cs-Earthworms only; 25P: 25mg/kg Cs-Plants
only; 25W+P: 25mg/kg Cs-Earthworms+Plants. 250N: 250mg/kg Cs-No biota; 250W: 250mg/kg CsEarthworms only; 250P: 250mg/kg Cs-Plants only; 250W+P: 250mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.2.4 Effects of Cs on earthworms
3.2.4.1 Earthworm survival
After 2 weeks, there was a significant difference in earthworm survival between
treatments (P=0.004) which were between 250 mg/kg compared with the control
(P=0.003) and the 25 mg/kg treatment (P=0.012). However, there was no significant
difference in earthworm survival between the earthworms only and earthworms+plants
ISMs (P=0.766) (Figure 3.3).
After 4 weeks, there was also a significant difference in earthworm survival between
treatments (P<<0.001) which were between 250 mg/kg compared with control
(P=0.001) and 25 mg/kg (P=0.001). Earthworms in 250 mg/kg had lower survival than
control and 25 mg/kg. However, there was no significant difference between survival
of earthworms in ISMs with earthworms only and with earthworms+plants (P=0.634)
(Figure 3.3).
55
Earthworm survival
2 weeks
120
4 weeks
Survival (%)
100
80
60
40
20
0
CW
CW+P
25W
25W+P
250W
250W+P
Treatments
Figure 3.3Earthworm survival after 2 and 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. 25W: 25mg/kg Cs-Earthworms only;
25W+P: 25mg/kg Cs-Earthworms+Plants; 250W: 250mg/kg Cs-Earthworms only; 250W+P: 250mg/kg CsEarthworms+Plants.
(The bars are means of 4 replicates ± standard error)
From 2 to 4 weeks, there was no significant difference in earthworm survival either in
control (P=0.170) or 25 mg/kg (P=0.142). However, there was a significant difference
in the 250 mg/kg treatment (P<<0.001). The earthworm survival in the 250 mg/kg
treatment was reduced after 4 weeks (Figure 3.3).
3.2.4.2 Effects of Cs on earthworm distribution
In the top layer soils:
After 2 weeks, there was a significant difference in earthworm distribution between
treatments (P=0.004) which were between control compared with 25 mg/kg (P=0.028)
and 250 mg/kg (P=0.001) treatments. The number of earthworms in the control was
higher than in 25 mg/kg and 250 mg/kg (5 ± 0.3, 3.6 ± 0.6, and 2.8 ± 0.5 respectively).
There was no significant difference in number of earthworms between the earthworms
only and earthworms+plants ISMs (P=0.6). There was no significant difference within
each treatment (P=0.061, P=0.303, and P=0.267 for control, 25 mg/kg and 250 mg/kg
treatments) (Figure 3.4).
56
After 4 weeks, there was no significant difference in earthworm distribution between
treatments (P=0.350) and between the ISMs with different biota across treatments
(P=0.553). There was only a significant difference within the 250 mg/kg (P=0.037). In
this treatment, the number of earthworms was found to be higher in the ISMs with
earthworms+plants (Figure 3.4)..
From 2 to 4 weeks, there was a significant decrease in earthworms in the control
(P=0.004). However, there were no significant changes in the 25 mg/kg (P=0.626) and
250 mg/kg (P=0.227) treatments (Figure 3.4).
120.0
100.0
80.0
60.0
40.0
Top layer
20.0
Bottom layer
2 weeks
250W+P
250W
25W+P
25W
CW+P
CW
250W+P
250W
25W+P
25W
CW+P
0.0
CW
The presence percentage of earthworms
Earthworm distribution
4 weeks
Treatments
Figure 3.4 Effects of Cs on earthworm distribution after 2 and 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. 25W: 25mg/kg Cs-Earthworms only;
25W+P: 25mg/kg Cs-Earthworms+Plants; 250W: 250mg/kg Cs-Earthworms only; 250W+P: 250mg/kg CsEarthworms+Plants.
(The bars are means of 4 replicates ± standard error)
57
In the bottom layer soils:
After 2 weeks, there was no significant difference between treatments (P=0.462) and
between treatments with different biota (P=0.327. There was also no significant
difference within control (P=0.537), 25 mg/kg (P=0.437), and 250 mg/kg (P=0.708)
(Figure 3.4).
After 4 weeks, there was no significant difference in earthworm distribution between
treatments (P=0.221), and between ISMs with different biota (P=0.681). There was
also no significant difference within each treatment (P=1.00, P=0.877, and P=0.457 for
control, 25 mg/kg and 250 mg/kg, respectively) (Figure 3.4).
From 2 to 4 weeks, there was no significant change in the number of earthworms in
control (P=0.402), 25 mg/kg (P=0.691), and 250 mg/kg (P=0.503) treatments (Figure
3.4).
Comparison of earthworm numbers in top and bottom soil layers:
After 2 weeks, there was no significant difference between earthworm numbers in top
and bottom soil in control (P=0.222), 25 mg/kg (P=0.106), and 250 mg/kg (P=0.208)
treatments (Figure 3.4).
However, after 4 weeks, a significant difference was observed in controls (P=0.040).
The number of earthworms in bottom soil layers in controls was higher than in the top
layer. There was no significant difference in the 25 mg/kg (P=0.135), and 250 mg/kg
treatments (P=0.197) (Figure 3.4).
3.2.4.3 Effects of Cs on earthworm biomass
After 2 weeks, there was a significant difference in earthworm biomass between
treatments (P=0.024) which were between the 250 mg/kg treatment compared with the
control (P=0.015) and the 25 mg/kg (P=0.046) treatment. Weight gain (%) of
earthworms from 250 mg/kg treatment was two times lower than in the control and the
25 mg/kg treatment. There was no significant difference between ISMs with different
58
biota across treatments (P=0.099). There was no significant difference within
treatments (P=0.559, P=0.468 and 0.306 for control, 25 and 250 mg/kg treatments
respectively) (Figure 3.5).
After 4 weeks, there was no significant difference between treatments (P=0.169),
between ISMs with different biota across treatments (P=0.487), and within treatments
(P=0.309, P=0.386, and P=0.058 for control, 25 mg/kg and 250 mg/kg treatments
respectively) (Figure 3.5).
From week 2 to week 4, the biomass of earthworms decreased significantly in the
control (P=0.025) and 250 mg/kg (P=0.027) but not in 25 mg/kg (P=0.051) (Figure
3.5).
Earthworm biomass gain
40.0
Biomass gain (%)
30.0
20.0
10.0
2 weeks
0.0
4 weeks
-10.0
-20.0
-30.0
CW
CW+P
25W
25W+P
250W
250W+P
Treatments
Figure 3.5 Effects of Cs on earthworm biomass
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. 25W: 25mg/kg Cs-Earthworms only;
25W+P: 25mg/kg Cs-Earthworms+Plants; 250W: 250mg/kg Cs-Earthworms only; 250W+P: 250mg/kg CsEarthworms+Plants.
(The bars are means of 4 replicates ± standard error)
59
3.2.4.4 Caesium accumulation in earthworms
Caesium in earthworms from top layer soils:
Caesium was detected in earthworms from the 25 mg/kg and 250 mg/kg treatments but
not in the controls. After 2 weeks, there was a significant difference between
treatments (P=0.001). Caesium concentration in earthworms from the 250 mg/kg
treatment was approximately seven times higher than in the earthworms from the 25
mg/kg treatment. There was no significant difference in Cs in earthworms between the
ISMs with different biota (earthworms only and earthworms+plants) across treatments
(P=0.753).
Within each treatment, there was no significant difference (P=0.386 and P=1.00 for 25
mg/kg and 250 mg/kg respectively) (Figure 3.6).
After 4 weeks, there was a significant difference in Cs accumulated by earthworms
between treatments (P=0.001). Cs was approximately 12 times higher in earthworms
from the 250 mg/kg treatment. However, there was no significant difference between
ISMs with different biota across treatments (P=0.248) and within the 25 mg/kg
treatment (P=0.386). However there was a significant difference within the 250 mg/kg
treatment with different biota (P=0.021). Earthworms from earthworms+plants ISMs
had higher Cs levels (309.8 ± 15.0 mg/kg) than in ISMs with only earthworms (229.7
± 15.2 mg/kg) (Figure 3.6).
From 2 to 4 weeks, there was no significant change in Cs concentration in earthworms
from the 25 mg/kg treatment (P=0.419). However, in the 250 mg/kg treatment,
earthworms accumulated more Cs at the second sampling (P=0.005) (178.1 ± 12.9
mg/kg and 269.7 ± 18.1 mg/kg for 2 and 4 weeks respectively) (Figure 3.6).
Caesium in earthworms from bottom layer soils:
After 2 weeks, there was a significant difference between treatments (P=0.001).
Caesium accumulated in earthworms from 250 mg/kg treatments was more than 13
60
times higher than earthworms from 25 mg/kg treatments. However, there was no
significant difference between earthworms from ISMs with different biota across
treatments (P=0.401), within 25 mg/kg (P=1.00) and within 250 mg/kg (P=0.564)
(Figure 3.6).
After 4 weeks, again there was a significant difference between treatments (P=0.001).
Earthworms from the higher 250 mg/kg Cs treatment accumulated more than 16 times
higher than those from the 25 mg/kg treatment. There was no significant difference
between Cs concentration in earthworms from ISMs with different biota across
treatments (P=0.916), within the 25 mg/kg (P=0.773), and within the 250 mg/kg
treatment (P=0.564) (Figure 3.6).
From 2 to 4 weeks, Cs concentration in earthworms from the 25 mg/kg treatment did
not change (P=0.310). However, in the 250 mg/kg treatment, there was a significant
increase of Cs in earthworms from 165.5 ± 12.4 mg/kg to 228.8 ± 17.5 mg/kg
(P=0.041) (Figure 3.6).
Comparison of Cs accumulation in earthworms from top and bottom soil layers:
After 2 weeks, in the 25 mg/kg treatment, Cs concentration in earthworms was more
twice that of earthworms in the top layer (P=0.004). Interestingly, a similar significant
difference was not observed in the 250 mg/kg treatment (P=0.485) (Figure 3.6).
After 4 weeks, there was no significant difference in Cs between earthworms in the top
and bottom layer in both the 25 mg/kg (P=0.106) and the 250 mg/kg treatments
(P=0.082) (Figure 3.6).
61
Cesium concentration (mg/kg)
Cesium concentration in earthworms and soils
400
Top soil
350
Top layer worms
300
Bottom soils
250
Bottom layer worms
200
150
100
50
0
CW
CW+P
25W
25W+P
250W
250W+P
CW
2 weeks
CW+P
25W
25W+P
250W
250W+P
4 weeks
Treatments
Figure 3.6 Concentration of Cs in earthworms and soil after 2 and 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. 25W: 25mg/kg Cs-Earthworms only; 25W+P: 25mg/kg Cs-Earthworms+Plants; 250W: 250mg/kg CsEarthworms only; 250W+P: 250mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
62
3.2.4.5 Caesium transfer factor (TF)
Transfer factor
Cesium transfer factor in earthworms (TF)
10
9
8
7
6
5
4
3
2
1
0
TFTop
TFBottom
25W
25W+P
250W
250W+P
25W
2 weeks
25W+P
250W
250W+P
4 weeks
Treatments
Figure 3.7 Caesium transfer factor in earthworms after 2 and 4 weeks
25W: 25mg/kg Cs-Earthworms only; 25W+P: 25mg/kg Cs-Earthworms+Plants. 250W: 250mg/kg CsEarthworms only; 250W+P: 250mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
TF for earthworms in top layer soils:
After 2 weeks, there was no significant difference in the Cs TF for earthworms
between treatments (P=0.462), and between ISMs with earthworms only and
earthworms+plants (biota) across treatments (P=0.505). There was no significant
difference detected within 25 mg/kg (P=0.886) and 250 mg/kg treatments (P=0.686)
(Figure 3.7).
After 4 weeks, there was also no significant difference in the Cs TF for earthworms
between treatments (P=0.172), and between ISMs with different biota across
treatments (P=0.294). In addition, there no significant difference within 25 mg/kg
(P=0.773), and within 250 mg/kg (P=0.248) treatments (Figure 3.7).
The results revealed that the Cs TF values for earthworms were higher after 4 weeks
(0.8 ± 0.06) compared with the TF after 2 weeks (0.6 ± 0.04) (P=0.026) (Figure 3.7).
63
Earthworms in bottom layer soils:
After 2 weeks, there was significant difference between treatments (P=0.006). TF
values for 250 mg/kg treatment earthworms were nearly double that of the lower
treatment. However, there was no significant difference between TF in earthworms in
ISMs with different biota across treatments (P=0.172). Hence, there was no significant
difference within 25 mg/kg (P=0.149), and within 250 mg/kg (P=0.248) treatments
(Figure 3.7).
After 4 weeks, there was also a significant difference between treatments (P=0.006).
TF values of earthworms from the 250 mg/kg treatments were more than double that
of earthworms from the 25 mg/kg treatments. There was no significant difference
between the Cs TF in earthworms from ISMs with different biota across treatments
(P=0.674). In addition, significant differences was not detected within 25 mg/kg
(P=0.564), and within 250 mg/kg (P=0.773) treatments (Figure 3.7).
Overall, the Cs TF values increased more than 1.5 times from 2 to 4 weeks (P=0.012).
Comparisons between earthworms in top and bottom soil layers:
After 2 weeks, in both 25 mg/kg and 250 mg/kg treatments, TF values were higher in
earthworms in bottom layer soils (P<0.001 for both cases). They were more than 3 and
7 times higher in the 25 mg/kg and 250 mg/kg treatments, respectively (Figure 3.7).
After 4 weeks, there was also a significant difference in TF values between
earthworms from top and bottom layers in both 25 mg/kg (P=0.003) and 250 mg/kg
(P=0.001) treatments. TF values were always higher in the earthworms from the
bottom layer earthworms compared with those from the top (more than 4 times in the
25 mg/kg treatment and nearly 8 times in the 250 mg/kg treatment) (Figure 3.7).
64
3.2.4.6 Total uptake of Cs by earthworms
Cesium total uptake in earthworms
Total cesium (mg)
40
35
Top Total Cs
30
Bottom Total Cs
25
20
15
10
5
0
25W
25W+P
250W 250W+P
25W
25W+P
2 weeks
250W 250W+P
4 weeks
Treatments
Figure 3.8 Total uptake of Cs by earthworms after 2 and 4 weeks
25W: 25mg/kg Cs-Earthworms only; 25W+P: 25mg/kg Cs-Earthworms+Plants. 250W: 250mg/kg CsEarthworms only; 250W+P: 250mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
Earthworms in top layer soils:
After 2 weeks, there was a significant difference between treatments (P=0.003) in total
uptake of Cs in earthworms. Earthworms in the 250 mg/kg treatment accumulated
nearly 5 times higher Cs than those in the 25 mg/kg treatment. There was no
significant difference between earthworms from ISMs with different biota across
treatments (P=0.833). In addition, there was no significant difference within 25 mg/kg
(P=0.144), and 250 mg/kg (P=0.083) treatments (Figure 3.8).
After 4 weeks, there was a significant difference in total Cs removed by earthworms,
between treatments (P=0.001). At this sampling time, earthworms from the 250 mg/kg
treatment accumulated more than 10 times higher Cs than those from the 25 mg/kg
treatment. There was no significant difference between earthworms from the ISMs
with different biota across treatments (P=0.752). There was no significant difference
65
within 25 mg/kg (P=0.139). However, there was a significant difference within the 250
mg/kg treatment (P=0.021). In this treatment, earthworms from the ISMs with
earthworms only accumulated nearly 3 times less than those from ISMs with
earthworms+plants (Figure 3.8).
From 2 to 4 weeks, the total uptake of Cs decreased nearly twice in the 25 mg/kg
treatment (P<<0.001). However, there was no significant difference in the 250 mg/kg
treatment (P=0.531) (Figure 3.8).
Earthworms in bottom layer soils:
After 2 weeks, there was a significant difference in total Cs uptake by earthworms
between treatments (P=0.001). Earthworms exposed to 250 mg/kg accumulated more
than 10 times higher Cs than earthworms exposed to 25 mg/kg. There was no
significant difference between total Cs in earthworms from ISMs with different biota
across treatments (P=0.713). Furthermore, there was no significant difference within
25 mg/kg (P=0.306), and within 250 mg/kg (P=1.00) treatments (Figure 3.8).
After 4 weeks, there was also a significant difference between treatments (P=0.001).
The accumulation by earthworms was more than 11 times higher in the 250 mg/kg
treatment. However, there was no significant difference between total Cs in
earthworms from ISMs with different biota across treatments (P=0.958). In addition, a
significant difference was not detected within 25 mg/kg (P=0.564), and 250 mg/kg
(P=0.663) treatments (Figure 3.8).
From 2 to 4 weeks, there was no significant changes in both 25 mg/kg (P=0.459) and
250 mg/kg (P=0.901) treatments in total uptake of Cs in earthworms (Figure 3.8).
Comparisons between earthworms in top and bottom layer soils:
After 2 weeks, in the 25 mg/kg treatment, there was a significant difference in Cs total
uptake between earthworms in the top and bottom layer soils (P=0.025). Earthworms
found in the top layer soils accumulated more Cs than the ones found in the bottom
66
soil layer. Conversely, the significant difference was not observed in 250 mg/kg
treatment (P=0.207) (Figure 3.8).
After 4 weeks, there was no significant difference in total uptake of Cs between
earthworms in the top and bottom layer soils in both 25 mg/kg (P=0.794) and 250
mg/kg (P=0.631) treatments (Figure 3.8).
3.2.5 Effects of Cs on plants
3.2.5.1 Effects of Cs on plant germination rate
There was no significant difference between treatments (P=0.647), between plant
germination rate in ISMs with different biota (plants only and earthworms+plants)
across treatments (P=0.304). Furthermore, there was no significant difference within
control (P=0.457), within 25 mg/kg (P=0.647) and within 250 mg/kg (P=0.119)
treatments (Figure 3.9).
Germination rate of plants
120
Germination rate (%)
100
80
60
40
20
0
CP
CW+P
25P
25W+P
250P
250W+P
Treatments
Figure 3.9 Effects of Cs on plant germination rate
CP: Control-Plants only; CW+P: Control-Earthworms+Plants. 25P: 25 mg/kg Cs-Plants only; 25W+P: 25 mg/kg
Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only; 250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
67
3.2.5.2 Effects of Cs on plant growth (shoot length)
After 2 weeks, there was no significant difference between plant shoot length in
treatments (P=0.057), and between plants from ISMs with different biota across
treatments (P=0.303). Hence, there was no significant difference within control
(P=0.073) and within 250 mg/kg treatments (P=0.225). However, there was a
significant difference within the 25 mg/kg treatment (P=0.012). In this treatment, the
length of plants from ISMs with earthworms+plants was lower than from those with
plants only (Figure 3.10).
After 4 weeks, there was no significant difference in plant shoot length between
treatments (P=0.756), and between plants from ISMs with different biota across
treatments (P=0.730). In addition, there was no significant difference within control
(P=0.707), within 25 mg/kg (P=0.583), and within 250 mg/kg (P=0.600) treatments
(Figure 3.10).
From week 2 to week 4, there was a significant increase in the shoot length between
plants from control (P=0.003), 25 mg/kg (P=0.004), but not 250 mg/kg treatments
(Figure 3.10).
Shoot length
2 weeks
30.0
4 weeks
Height (cm)
25.0
20.0
15.0
10.0
5.0
0.0
CP
CW+P
25P
25W+P
250P
250W+P
Treatments
Figure 3.10 Effects of Cs on shoot length after 2 and 4 weeks
CP: Control-Plants only; CW+P: Control-Earthworms+Plants. 25P: 25 mg/kgCs-Plants only; 25W+P: 25 mg/kg
Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only; 250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
68
3.2.5.3 Effects of Cs on plant biomass
After 2 weeks, there was a significant difference between treatments (P<0.001). Dried
biomass increased in order of control < 25 mg/kg < 250 mg/kg. There was no
significant difference between plants from ISMs with different biota across treatments
(P=0.986). Furthermore, there was no significant difference within control (P=0.238),
25 mg/kg (P=0.471) and 250 mg/kg (P=0.225) treatments (Figure 3.11).
After 4 weeks, there was also a significant difference between treatments (P<0.001).
At this sampling time, plants from the 250 mg/kg treatment had the highest biomass
compared with control and those from the 25 mg/kg treatment. The order of biomass
remained similar to the biomass after 2 weeks. There was no significant difference
between plant biomass from ISMs with different biota across treatments (P=0.541);
within control (P=0.816), 25 mg/kg and 250 mg/kg treatments (P=0.382 and P=6.00
respectively) (Figure 3.11).
From 2 to 4 weeks, plant above ground dried biomass significantly increased in
Biomass (gram)
control, 25 mg/kg and 250 mg/kg treatments (P<0.001 for all cases) (Figure 3.11).
0.2
0.18
0.16
0.14
0.12
0.1
0.08
0.06
0.04
0.02
0
Plant above ground dried biomass
2 weeks
4 weeks
CP
CW+P
25P
25W+P
250P
250W+P
Treatments
Figure 3.11 Effects of Cs on above ground dried biomass after 2 and 4 weeks
CP: Control-Plants only; CW+P: Control-Earthworms+Plants. 25P: 25 mg/kgCs-Plants only; 25W+P: 25 mg/kg
Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only; 250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
69
3.2.5.4 Caesium accumulation by plants
Caesium was not detected in shoots and roots of plants in controls. Caesium was
detected in plants from both 25 mg/kg and 250 mg/kg treatments (Figure 3.12).
Caesium in shoots:
After 2 weeks, there was a significant difference between treatments (P<0.001).
Caesium concentration in shoots was more than 17 times higher in the 250 mg/kg
treatment compared to plant shoots in the 25 mg/kg treatment. There was no
significant difference between biota across treatments (P=1.00); within 25 mg/kg
(P=0.115), and 250 mg/kg (P=0.115) treatments (Figure 3.12).
After 4 weeks, there was a significant difference between treatments (P<<0.001). At
this sampling time, Cs in 250 mg/kg treatment plant shoots was more than 13 times
higher than in plants from 25 mg/kg treatment. There was no significant difference
between plants from ISMs with different biota across treatments (P=0.880); within 25
mg/kg (P=0.208), and within 250 mg/kg (P=0.093) treatments (Figure 3.12).
From 2 to 4 weeks, there was no significant change in Cs concentration in plants from
both 25 mg/kg (P=0.969) and 250 mg/kg (P=0.114) treatments (Figure 3.12).
Cs in plant roots:
After 2 weeks, Cs concentration in 250 mg/kg treatment plant roots was more than 15
times higher than 25 mg/kg plants (P=0.001). However, there was no significant
difference between plants from ISMs with different biota (plants only and
earthworms+plants) across treatments (P=1.00); within 25 mg/kg (P=0.564); and
within 250 mg/kg (P=0.564) (Figure 3.12).
70
After 4 weeks, again Cs concentration in plant roots from the 250 mg/kg treatment was
more than 5 times higher than in those from the 25 mg/kg treatment (P=0.001). There
was no significant difference between plants from ISMs with different biota across
treatments (P=0.674), and within 250 mg/kg treatments (P=0.248). However, there
was significant difference within 25 mg/kg treatments (P=0.021). Caesium
concentration of plants in the plants only ISMs (227.5 ± 24.3 mg/kg) was higher than
in plants from earthworms+plants ISMs (147.7 ± 9.8 mg/kg) (Figure 3.12).
From 2 to 4 weeks, Cs level was not significantly altered in plant roots from the 25
mg/kg treatment plants (P=0.580). However, Cs concentration was significantly
reduced from 3322.4 ±328.6 mg/kg to 1042.2 ± 93.0 mg/kg in the plants from the 250
mg/kg treatment (P=0.001) (Figure3.12).
71
Cesium concentration in plants
4500
4000
Top soils Cs
Cesium concentration (mg/kg)
3500
Shoots Cs
Roots Cs
3000
2500
2000
1500
1000
500
0
CP
CW+P
25P
25W+P
250P
250W+P
CP
2 weeks
CW+P
25P
25W+P
250P
250W+P
4 weeks
Treatments
Figure 3.12 Concentration of Cs in plants after 2 and 4 weeks
CP: Control-Plants only; CW+P: Control-Earthworms+Plants. 25P: 25 mg/kg Cs-Plants only; 25W+P: 25 mg/kg Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only; 250W+P:
250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
72
3.2.5.5 Caesium transfer factor (TF) in plants
After 2 weeks, the TF values for plants were more than 2 times higher in 250 mg/kg
compared with 25 mg/kg treatment (P=0.001). The presence of earthworms did not
increase the TF values across treatments (P=0.734). There was no significant
difference within 25 mg/kg (P=0.074), and within 250 mg/kg (P=0.248) treatments
(Figure 3.13).
After 4 weeks, interestingly, there was no significant difference in TF values for Cs in
plants between treatments (P=0.258); between plants from ISMs with different biota
across treatments (P=0.910); within 25 mg/kg (P=0.294), and within 250 mg/kg
treatments (P=0.074)(Figure 3.13).
From 2 to 4 weeks, Cs TF values in plants were not significantly changed in both 25
mg/kg (P=0.569) and 250 mg/kg (P=0.06) treatments (Figure 3.13).
Transfer factor of cesium in sunflowers (TF)
6
Transfer factor
5
4
3
2
1
0
25P
25W+P
250P
250W+P
25P
2 weeks
25W+P
250P
250W+P
4 weeks
Treatments
Figure 3.13.Caesium transfer factor in plants after 2 and 4 weeks
25P: 25 mg/kg Cs-Plants only; 25W+P: 25 mg/kg Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only;
250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
73
3.2.5.6 Comparison of shoot and roots -translocation factor
After 2 weeks, there was no significant difference in translocation factor between
treatments (P=0.600) and between plants from ISMs with different biota across
treatments (P=0.916). In addition, there was no significant difference within 25 mg/kg
(P=0.248), and within 250 mg/kg (P=0.248) treatments (Figure 3.14).
At the second sampling time, there was also no significant difference between
treatments (P=0.093), between plants from ISMs with different biota across treatments
(P=0.115), within 25 mg/kg (P=0.149), and within 250 mg/kg (P=0.386) treatments
(Figure 3.14).
From the first to the second sampling, there was no significant change in translocation
factor of plants from both 25 mg/kg and 250 mg/kg treatments (P=0.913, and P=0.144
respectively) (Figure 3.14).
Cesium translocation factor in sunflower
Translocation factor
1
0.8
0.6
0.4
0.2
0
25P
25W+P
250P
250W+P
25P
2 weeks
25W+P
250P
250W+P
4 weeks
Treatments
Figure 3.14Caesium translocation factor in plants after 2 and 4 weeks
25P: 25 mg/kg Cs-Plants only; 25W+P: 25 mg/kg Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only;
250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
74
3.2.5.7 Caesium total uptake in plants
After 2 weeks, there was a significant difference between treatments (P<<0.001).
Plants in 250 mg/kg accumulated much higher amounts of Cs compared with 25
mg/kg treatment plants (nearly 22 times higher). The presence of earthworms did not
affect Cs total uptake across treatments (P=0.651), within 25 mg/kg (P=0.600), and
within 250 mg/kg (P=0.462) treatments (Figure 3.15).
After 4 weeks, there were slight changes. Plants in 250 mg/kg treatment extracted
higher Cs than 25 mg/kg plants (nearly 16 times) (P<<0.001). There was no significant
difference between plants from ISMs with different biota across treatments (P=0.851)
and within the 250 mg/kg treatment (P=0.115). However, within the 25 mg/kg
treatment, Cs total uptake from earthworms+plants was lower than that from plants
only (P=0.036) (Figure 3.15).
From 2 to 4 weeks, there was a significant increase in both 25 mg/kg (P=0.006) and
250 mg/kg (P=0.007) treatments in Cs total uptake by sunflowers. The increase in total
uptake of Cs was nearly 2.5 times for plants from the 25 mg/kg treatment in this period
and 1.8 times for those from the 250 mg/kg treatment (Figure 3.15).
Cesium total uptake in sunflowers
Total cesium (mg)
120
100
80
60
40
20
0
25P
25W+P
250P
250W+P
25P
2 weeks
25W+P
250P
250W+P
4 weeks
Treatments
Figure 3.15 Caesium total uptake in plants after 2 and 4 weeks
25P: 25 mg/kg Cs-Plants only; 25W+P: 25 mg/kg Cs-Earthworms+Plants; 250P: 250 mg/kg Cs-Plants only;
250W+P: 250 mg/kg Cs-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
75
3.3 Lead and Caesium+Lead as a mixture experiment
3.3.1 Soil parameters after the 4 week experiment
Results revealed that all pH values were ranged from 6.4 - 7.0 in top layer soils and
from 6.3 – 7.0 in bottom layer soils (Table 3.7). Soil conductivity was higher in the
bottom layer soils and in contaminated soils. However, these values were in the range
from 0 to 2000 μS/cm which was considered to be the conductivity range that causes
no adverse effects on plants (Table 3.8).Soil moisture was higher in the bottom layer
soils and ranged from 25% to 36% in top layer soils and from 30% to 40% (Table 3.9).
Organic matter content did not change after 4 weeks and remained between 11-12 %
(Table 3.10).
Table 3.7 Soil pH after 4 weeks (mean ± SE, n=3)
layers
No biota
Earthworms only
Plants only
Worms+Plants
Top
6.9±0.04
7.0±0.01
7.0±0.03
6.9±0.05
Bottom
6.9±0.02
6.9±0.02
7.0±0.05
6.9±0.02
Top
6.5±0.00
6.6±0.02
6.4±0.00
6.5±0.03
Bottom
6.4±0.00
6.5±0.01
6.4±0.03
6.5±0.00
Top
6.5±0.01
6.4±0.00
6.4±0.04
6.4±0.03
Bottom
6.6±0.02
6.5±0.01
6.4±0.01
6.4±0.03
Control
Pb only
Cs+Pb
Biota
Soil
Treatment
Table 3.8 Soil conductivity (μS/cm) after 4 weeks (mean ± SE, n=3)
Treatment
Control
Pb only
Cs+Pb
Biota
Soil
layers
No biota
Earthworms only
Plants only
Worms+Plants
Top
62.23±1.4
63.00±3.3
50.03±0.95
45.30±0.40
Bottom
83.63±2.2
74.00±0.5
54.23±2.40
49.27±1.51
Top
90.47±2.15
56.27±3.58
90.60±2.56
97.13±0.46
Bottom
138.60±12.88
84.30±1.65
174.30±8.17
123.43±5.69
Top
72.27±5.26
76.07±1.50
75.13±2.37
85.57±0.57
Bottom
91.63±0.74
85.40±4.16
142.07±2.89
149.47±26.33
76
Table 3.9 Soil moisture (%) after 4 weeks(mean ± SE, n=3)
Treatment
Control
Pb only
Cs+Pb
Biota
Soil
layers
No biota
Earthworms only
Plants only
Worms+Plants
Top
25.22±0.25
25.34±0.28
26.47±0.59
28.23±0.38
Bottom
34.19±1.00
32.26±0.13
34.41±0.25
31.07±0.26
Top
30.52±0.26
31.61±0.31
32.33±0.17
30.50±1.01
Bottom
38.06±0.35
36.19±0.10
38.06±0.03
36.30±0.34
Top
32.34±0.53
36.05±0.13
33.09±0.20
32.23±0.19
Bottom
39.23±0.57
40.05±0.93
40.21±0.12
37.53±0.30
Table 3.10 Soil organic matter content (%) after 4 weeks (mean ± SE, n=3)
Treatments
Control
Pb only
Cs+Pb
Soil
layers
Biota
No biota
Earthworms
Plants only
Worms+Plants
only
Top
12.41±0.11
12.44±0.06
12.28±0.08
12.02±0.07
Bottom
12.34±0.04
12.26±0.14
12.18±0.05
11.91±0.14
Top
11.75±0.25
11.47±0.13
11.47±0.03
11.24±0.03
Bottom
11.74±0.40
11.87±0.07
11.42±0.02
11.45±0.24
Top
11.05±0.03
10.89±0.02
10.91±0.16
10.99±0.17
Bottom
11.37±0.11
10.92±0.13
10.96±0.09
10.91±0.08
3.3.2 Lead concentration in soils
To facilitate data analysis in this experiment the data was separated into top and
bottom soil layers and the control was removed from the data list. The two way
ANOVA was then used. In the top soil layers, there was no significant difference
between treatments (Pb only and Cs+Pb) (P=0.570) and between soil from ISMs with
different biota (P=0.122) (Figure 3.16).
Pb was detected in the bottom soil layers (Fig 3.16). The data was tested for normality
and since the variance was not equal between means, the Kruskal-Wallis test was used
to test the difference between treatments. No significant difference between treatments
was found (P=0.079). Within the Pb only treatment, there was a significant difference
between soil from ISMs with different biota (P=0.001). In the Pb only treatment, the
Pb concentration in the bottom layer with the presence of earthworms alone was
significantly higher (183 ± 6.5 mg/kg) when compared with the soil from ISMs with
77
no biota, earthworms+plants, and plants only (90.4 ± 9.1, 130.3 ± 17.2, and 134.3 ±
9.9 mg/kg respectively). However, the difference was not detected in the Cs+Pb
treatment (P=0.904) (Figure 3.16).
Overall, there was a significant difference between Pb in top and bottom soil layers
(P=0.046). Lead remained mostly in the top (1,419 ± 25.5 mg/kg) and not in the
bottom (139 ± 7.4 mg/kg) of the soil core (Figure 3.16).
Lead concentration in soils
2000
Lead concentration (mg/kg)
1800
Top layer
Bottom layer
1600
1400
1200
1000
800
600
400
200
0
IL
ILC
CN
CW
CP
CW+P
LN
LW
LP
LW+P
LCN
LCW
LCP
LCW+P
Treatments
Figure 3.16 Lead concentrations in soils after 4 weeks
IL: Initial soil-Pb only; ILC-Initial soil-Cs+Pb. CN: Control-No biota;CW: Control-Earthworms only; CP:
Control-Plants only; CW+P: Control-Earthworms+Plants. LN: Pb-No biota;LW: Pb-Earthworms only; LP: PbPlants only; LW+P: Pb-Earthworms+Plants. LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP:
Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.3 Caesium concentration in soils
Caesium was measured only in the Cs+Pb treatment. In both top and bottom layer
soils, there was no significant difference between soil from ISMs with different biota
(P=0.693, and P=0.827 respectively). The results revealed that Cs moved down
through soil core. There was a significant difference in concentration of Cs between
top and bottom layers of soils (P<<0.05). Caesium was higher in the top compared
with bottom layer soils (Figure 3.17).
78
Top layer
Bottom layer
Concentration of Cesium (mg/ kg)
Cesium in soil layers
500
400
300
200
100
0
LCN
LCW
LCP
LCW+P
Treatments
Figure 3.17Caesium in soils after 4 weeks
LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP: Cs+Pb-Plants only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.3.4 Potassium concentration in soils
Potassium concentration (mg/ kg)
Potassium concentration in Soils
3500
3000
Top layer
2500
Bottom layer
2000
1500
1000
500
0
LCN
LCW
LCP
LCW+P
Biotas
Figure 3.18 Potassium concentrations in soils after 4 weeks
LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP: Cs+Pb-Plants only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 3 replicates ± standard error)
Potassium was also measured in the Cs+Pb treatment only. In both top and bottom soil
layers, there was no significant difference between soil from ISMs with different biota
79
(P=0.335, and P=0.912 respectively). In addition, the difference between K
concentrations in top and bottom soil layers was not significant (P=0.7) (Figure 3.18).
3.3.5 Effects of Pb only and Cs+Pb on leachates
3.3.5.1 pH of leachates
Table 3.11 Leachate pH after the four week experiment (mean ± SE, n=4)
Biota
Treatment
No biota
Earthworms
Plants only
Worms+Plants
only
Control
5.9±0.07
6.3±0.00
6.2±0.03
6.0±0.11
Pb only
5.9±0.06
6.2±0.06
5.9±0.11
5.7±0.11
Cs+Pb
6.4±0.04
6.3±0.12
5.8±0.01
5.8±0.05
Data was analysed using the Kruskal-Wallis test since the assumptions for ANOVA
could not be met. There was no significant difference between treatments (P=0.110),
Within the control, there was a significant difference in the pH of leachates from ISMs
with different biota (P=0.003,). Without biota and with earthworms+plants, the pH of
leachates was significantly lower than in ISMs with earthworms only (P=0.006 and
P=0.014 respectively) (Table 3.11).
Within Pb only treatments, there was also significant difference in the pH of leachates
from ISMs with different biota (P=0.013, one way ANOVA was used) which was
between the ISMs with earthworms only and earthworms+plants (P=0.008) and the pH
of leachates from the earthworms+plants ISMs was lower than those from the ISMs
with earthworms only (Table 3.11).
Within Cs+Pb, significant differences in the pH of leachates from ISMs with different
biota was observed (P=0.009, Kruskal-Wallis test). The pH of leachates from
treatments without biota was significantly higher than with plants only and
earthworms+plants (P<< 0.001 for both). The leachate pH was higher with earthworms
only than with plants only (P=0.008) and earthworms+plants (P=0.006) (Table 3.11).
80
3.3.5.2 Lead concentration in leachates
Lead concentration in Leachates
Lead concentration (mg/L)
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0.00
CN
CW
CP
CW+P
LN
LW
LP
LW+P
LCN
LCW
LCP LCW+P
Treatments
Figure 3.19 Lead concentrations in leachates after 4 weeks
CN: Control-No biota;CW: Control-Earthworms only; CP: Control-Plants only; CW+P: ControlEarthworms+Plants. LN: Pb-No biota;LW: Pb-Earthworms only; LP: Pb-Plants only; LW+P: PbEarthworms+Plants. LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP: Cs+Pb-Plants only; LCW+P:
Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
Lead was detected in the Pb only and Cs+Pb leachates but not in controls after 4
weeks. There was a significant difference between treatments (Pb only and Cs+Pb)
(P=0.001). Lead was higher in Pb only treatment leachates compared with Cs+Pb
treatment leachates (0.055±0.012 mg/L, and 0.034 ±0.003 mg/L respectively). There
was also a significant difference in the pH of leachates from ISMs with different biota
(P<<0.05). Leachates from ISMs without biota had the lowest Pb concentration (0.02
± 0.01 mg/L). Leachates from ISMs with earthworms only (0.05 ± 0.01 mg/L)
contained less Pb than those with plants only (0.07 ± 0.01 mg/L). Leachates from
ISMs with earthworms+plants (0.04 ± 0.01 mg/L) had lower Pb than leachates form
ISMs with plants only (Figure 3.19).
Within the Pb single metal treatment, the no Pb was measured in leachates from ISMs
without biota (0 mg/L). The highest concentration of Pb was observed in leachates
from ISMs with plants only (0.12 ± 0.05 mg/L). Within the Cs+Pb treatment , the
81
highest Pb concentration was in leachates from ISMs with earthworms+plants (0.045 ±
0.001 mg/L) and the lowest was from ISMs with plants only (0.02 ± 0.001 mg/L)
(Figure 3.19).
3.3.5.3 Caesium and potassium concentration in leachates
Cesium concentration (mg/ L)
Cesium concentration in leachates.
0.3
0.25
0.2
0.15
0.1
0.05
0
LCN
LCW
LCP
LCW+P
Treatments
Figure 3.20 Caesiumconcentration in leachates after 4 weeks
LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP: Cs+Pb-Plants only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 3 replicates ± standard error)
After 4 weeks, Cs was only detected in leachates in the Cs+Pb treatment. The KruskalWallis test was used due to unequal variances in the data. There was no effect of biota
on leachate Cs concentration (P=0.053) (Figure 3.20).
However, the concentration of potassium in leachates were significantly different
between ISMs with different biota (P=0.001). Leachates from ISMs with
earthworms+plants (43.7±2.6 mg/L) had significantly higher potassium concentration
when compared with ISMs without biota (22.0± 1.0 mg/L) (P=0.001), earthworms
only (23.9±4.1 mg/L) (P=0.002), and plants only (26.9±1.0 mg/L) (P=0.007) (Figure
3.21).
82
Potassium concentration (mg/ L)
Potassium concentration in leachates.
50
45
40
35
30
25
20
15
10
5
0
LCN
LCW
LCP
LCW+P
Biotas
Figure 3.21 Potassium concentrations in leachates after 4 weeks
LCN: Cs+Pb-No biota; LCW: Cs+Pb-Earthworms only; LCP: Cs+Pb-Plants only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.3.6 Effects of Pb only and Cs+Pb as a mixture on earthworms
3.3.6.1 Effects of Pb only and Cs+Pb on earthworm survival
Earthworm survival
Survival percentage (%)
120
100
80
60
40
20
0
CW
CW+P
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.22 Effects of Pb and Cs+Pb on earthworm survival after 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. LW: Pb-Earthworms only; LW+P: PbEarthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
83
There was no significant difference in earthworm survival between treatments
(P=0.246). Hence, there was no significant difference between ISMs with different
biota across treatments (P=0.428). There was also no significant difference within
controls (P=0.495), Pb only (P=0.765), and Cs+Pb (P=0.369) (Figure 3.22).
3.3.6.2 Effects of Pb only and Cs+Pb on earthworm distribution
In top layer soils, the number of earthworms were not significantly different between
treatments (P=0.130) and between ISMs with different biota across treatments
(P=0.165). There was no significant difference within control (P=0.405), Pb only
(P=0.102), and Cs+Pb (P=1.0) treatments (Figure 3.23).
In bottom layer soils, there was a significant difference in the number of earthworms
between treatments (P=0.006). The number of earthworms in the bottom soil layer was
higher in controls compared with Pb only (P=0.048) and Cs+Pb (P=0.002) treatments.
However, there was no significant difference between the number of earthworms in
the bottom soil layer in Pb only and Cs+Pb treatments (P=0.304). There was no
significant difference between ISMs with different biota (earthworms only and
earthworms+plants) across treatments (P=0.258). Within each treatment, there was a
significant difference in the Pb single metal treatment: the number of earthworms in
the bottom layer from ISMs with earthworms+plants was higher than in those with
earthworms only (P=0.013) (Figure 3.23).
When the number of earthworms in top and bottom layers, in controls and Pb only
treatments were compared there were no significant difference (P=0.442, and P=0.291
respectively). However, there was a significant difference in Cs+Pb treatments
(P=0.002) with a higher number of earthworms in the top layer (P=0.03) (Figure 3.23).
84
Earthworm presence percentage (%)
Earthworm distribution
120%
100%
Top layer
Bottom layer
80%
60%
40%
20%
0%
CW
CW+P
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.23 Effects of Pb and Cs+Pb on earthworm distribution after 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. LW: Pb-Earthworms only; LW+P: PbEarthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.6.3 Effects of Pb only and Cs+Pb on earthworm fresh biomass
Percentage of weight loss (%)
Percentage of weight loss
25
20
15
10
5
0
CW
CW+P
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.24 Effects of Pb only and Cs+Pb on earthworm biomass after 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. LW: Pb-Earthworms only; LW+P: PbEarthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
85
Effects of Pb alone and Cs+Pb on earthworm biomass were based on percentage of
weight loss which was calculated by the difference between the initial and final fresh
weight of earthworms. There was a significant difference between treatments
(P<<0.001) and there was a significant effect of the different biota (P=0.049) (Figure
3.24).
The weight loss of control earthworms was less than in the Pb only treatment
(P<<0.001) and with Cs+Pb (P=0.029). The weight loss of earthworms was higher in
ISMs with earthworms+plants (17.1 ± 1.2%) compared with those with earthworms
only (14.3 ± 1.5%) (Figure 3.24).
3.3.6.4 Lead accumulation in earthworms
Lead concentration in soil and earthworms
Lead concentration (mg/kg)
1600
1400
1200
1000
800
Top layer soils
Top layer worms
Bottom layer soils
Bottom layer worms
600
400
200
0
CW
CW+P
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.25 Lead concentration in soil and earthworms after 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. LW: Pb-Earthworms only; LW+P: PbEarthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
Earthworms accumulated Pb in all treatments. Earthworms from the Pb only treatment
and Cs+Pb had very much higher Pb concentrations compared with controls (Figure
3.25). In top earthworms from the top soil layer, Pb concentration in the Pb only
treatment was significant higher than in the Cs+Pb treatment (P=0.025) (372.22 ± 9.13
mg/kg, 326.30 ± 15.94 mg/kg respectively). There was no significant difference
86
between earthworms from ISMs with different biota across treatments (P=0.529).
Within each treatment, there was no significant effect of biota (earthworms only and
earthworms+plants) (P=0.273 and P=0.773 for Pb only and Cs+Pb respectively)
(Figure 3.25).
In earthworms from the bottom soil layer, there was no significant difference in Pb
concentration between treatments (P=0.91). However, there was a significant effect of
biota across treatments (P=0.46). Lead concentration of earthworms in ISMs with only
earthworms was higher than in those from ISMs with earthworms+plants regardless of
which treatment they were from (302.32 ± 21.78 mg/kg, 244.07 ± 10.61 mg/kg). There
was no significant differences within the Pb only treatment (P=0.386). However, there
was a significant difference within the Cs+Pb treatment (P=0.021). Lead concentration
in ISMs with earthworms only was higher than earthworms+plants (303.75 ± 17.5
mg/kg and 245.67 ± 16.3 mg/kg respectively) (Figure 3.25).
There was a significant difference in Pb cocentration in earthworms in the top layer
and bottom layer (P=0.001). Lead concentration in earthworms was higher in the top
layer (349.26 ± 10.67 mg/kg) compared with those in the bottom layer (273.20 ± 13.91
mg/kg) (Figure 3.25).
3.3.6.5 Lead transfer factor in earthworms
In top layer soil earthworms, there was a significant difference between treatments
(P=0.035). TF values in earthworms from Pb single metal exposures (0.28 ± 0.01)
were higher than in those with Cs+Pb (0.24 ± 0.01) ones. However, there was no
significant effect of biota on TF of Pb in earthworms across treatments (P=0.973). In
addition, within Pb only and Cs+Pb treatments there was also no significant difference
(P=0.388 and P=0.430 respectively) (Figure 3.26).
In earthworms from the bottom layer soil there was no significant difference between
treatments (P=0.423); and no effect of biota across treatments (P=0.8360). Within Pb
only and Cs+Pb treatment, there was no significant difference in Pb TF values for
earthworms (P=0.773 for both) (Figure 3.26).
87
lead transfer factor (TF)in earthworms
3
Transfer factor (TF)
2.5
2
1.5
Top layer worms
1
Bottom layer worms
0.5
0
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.26 Lead transfer factor in earthworms after 4 weeks
LW: Pb-Earthworms only; LW+P: Pb-Earthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.6.6 Lead total uptake in earthworms
Total lead (mg)
Pb total uptake in earthworms
16
Top layer worms
14
Bottom layer worms
12
10
8
6
4
2
0
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.27 Lead total uptake in earthworms after 4 weeks
LW: Pb-Earthworms only; LW+P: Pb-Earthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+PbEarthworms+Plants.
(The bars are means of 4 replicates ± standard error)
88
In earthworms from top layer soils, there was no significant difference in total Pb
uptake between treatments (P=0.225), and between earthworms from ISMs with
different biota across treatments (P=0.831). Within each treatment there was also no
significant difference (P=0.856 and P=0.891 for Pb only and Cs+Pb respectively)
(Figure 3.27).
In earthworms from the bottom layer there was no significant difference between
treatments (P=0.507), and no effect of biota across treatments (P=0.507). Within each
treatment, there was no significant difference (P=0.562 and P=0.765 for Pb only and
Cs+Pb respectively) (Figure 3.27).
Comparison of earthworms from top and bottom soil layers:
In the Pb single metal treatment, Pb total uptake was higher in earthworms from the
top layer (P=0.011). The amount of Pb was 13.4 ± 0.4 mg and 9.3 ± 1.1 mg for
earthworms from the top and bottom layers respectively (Figure 3.27).
In the Cs+Pb treatment, a significant difference was not observed for total Pb uptake
between top and bottom earthworms (P=0.120) (Figure 3.27).
3.3.6.7 Caesium in earthworms
Cesium concentration (mg/kg)
Cesium concentration in soil and
earthworms
Top soil
Top layer worms
Bottom soil
Bottom layer worms
450
400
350
300
250
200
150
100
50
0
LCW
LCW+P
Biotas
Figure 3.28 Caesium concentration in earthworms after 4 weeks
LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
89
Caesium concentration was only tested in Cs+Pb treatments. The results revealed that
earthworms accumulated Cs. Caesium in earthworms from this treatment was higher
than in the control. Cs concentration in earthworms in both top and bottom layers was
not significantly different in ISMs with different biota (earthworms only and
earthworms+plants) (P=0.245 for top layer and P=0.713 for bottom layer soil
earthworms) (Figure 3.28).
3.3.6.8 Caesium transfer factor in earthworms
There was no significant difference of biota on the Cs transfer factor (TF) in
earthworms from either the top (P=0.373) or bottom layer of soil (P=0.783). However,
there was a significant difference in Cs TF values between earthworms in top and
bottom layer soil across the Cs+Pb treatment (P=0.001). The TF values of earthworms
in bottom layers (3.76 ± 0.5) were higher than those in the top layers (0.28 ± 0.03)
(Figure 3.29).
Transfer factors
Cesium transfer factor in earthworms (TF)
5
4.5
4
3.5
3
2.5
2
1.5
1
0.5
0
Worm top layer
Worm bottom layer
LCW
LCW+P
Biota
Figure 3.29Caesium transfer factor of earthworms after 4 weeks
LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants
(The bars are means of 3 replicates ± standard error)
90
3.3.6.9 Caesium total uptake in earthworms
There was no significant difference in total uptake of Cs between earthworms in the
ISMs with only earthworms and earthworms+plants (P=0.481). In addition, there was
also no significant difference in total uptake of Cs in earthworms from top and bottom
layer soils (P=0.969) (Figure 3.30).
Total cesium (mg)
Total cesium uptake in earthworms
5
4.5
4
3.5
3
2.5
2
1.5
1
0.5
0
TopWorm
BottomWorm
LCW
LCW+P
Biotas
Figure 3.30 Caesium total uptake in earthworms after 4 weeks
LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants
(The bars are means of 3 replicates ± standard error)
3.3.6.10 Potassium in earthworms
In earthworms from the top layer, there was significant difference between treatments
in potassium concentration (P=0.04). Potassium concentration of earthworms from the
Cs+Pb treatment (9,033.61 ± 219.66 mg/kg) was lower than that of control (10,906.40
± 400.92 mg/kg) and Pb only (10,232.45 ± 360.92 mg/kg) treatments. There was
however, no similar significant difference between control and Pb only treatments
(P=0.162). There was also no significant effect of biota (earthworms only and
earthworms+plants) across treatments (P=0.673) (Figure 3.31).
In earthworms from the bottom layer, the data was transformed (using 1/X) to achieve
a normal distribution. The analysis revealed that there was a significant difference
between treatments (P<<0.05). Potassium in earthworms from the Cs+Pb treatment
was significantly different from those in controls (P<<0.001) and Pb only (P<<0.001)
91
treatments. Meanwhile, K in earthworms from control and Pb only treatments was not
different (P=0.081). Again, K concentrations in earthworms from Cs+Pb treatments
(8,735.63 ± 177.03 mg/kg) was lower than in control (11,567.79 ± 601.52 mg/kg) and
Pb only (10,719.95 ± 172.70 mg/kg) treatments (Figure 3.31).
There was no significant difference in K levels in earthworms from top and bottom
layer soils (P=0.3) (Figure 3.31).
Potassium concentration (mg/ kg)
Potassium concentration in earthworms
16000
Top layer
14000
Bottom layer
12000
10000
8000
6000
4000
2000
0
CW
CW+P
LW
LW+P
LCW
LCW+P
Treatments
Figure 3.31 Potassium concentrations in earthworms after 4 weeks
CW: Control-Earthworms only; CW+P: Control-Earthworms+Plants. LW: Pb-Earthworms only; LW+P: PbEarthworms+Plants. LCW: Cs+Pb-Earthworms only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.3.7 Effects of Pb only and Cs+Pb as a mixture on plants
3.3.7.1 Plant germination
Germination rate of plants was not affected by chemicals (Pb or Cs+Pb) (P=0.624) and
biota (plants only or earthworms+plants) (P=0.169). Within each treatment, significant
differences were not detected (P=0.877, P=0.508, and P=0.063 for control, Pb only
and Cs+Pb treatments) (Figure 3.32).
92
Plant germination rate
Germination rate (%)
120.0
100.0
80.0
60.0
40.0
20.0
0.0
CP
CW+P
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.32 Effects of Pb only and Cs+Pb on plant germination rate
CP:
Control-Plants
only;
CW+P:
Control-Earthworms+Plants.
LP:
Pb-Plants
only;
LW+P:
Pb-
Earthworms+Plants. LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.7.2 Plant growth (shoot length)
Plant growth
14.0
Plant height (cm)
12.0
10.0
8.0
6.0
4.0
2.0
0.0
CP
CW+P
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.33 Effects of Pb only and Cs+Pb on plant growth (shoot length) after 4
weeks
CP:
Control-Plants
only;
CW+P:
Control-Earthworms+Plants.
LP:
Pb-Plants
only;
LW+P:
Pb-
Earthworms+Plants. LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
93
There was no significant difference between treatments (P=0.285) and no effect of
biota (P=0.685) on shoot length of plants. Within each treatment, there was no effect
of biota on plant growth (P=0.875, P=0.958, P=1.00 for control, Pb only and Cs+Pb
treatments) (Figure 3.33).
3.3.7.3 Above ground dried biomass of plants
There was no significant difference in dried biomass of plants between treatments
(P=0.099), between Biota (P=0.775) and there was no interaction between treatment
and biota in effects on plant dried biomass (P=0.694) (Figure 3.34).
Plant above ground dried Biomass
Dried weight/ plant (g)
0.25
0.20
0.15
0.10
0.05
0.00
CP
CW+P
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.34 Effects of Pb only and Cs+Pb on plant biomass (above ground dried
biomass) after 4 weeks
CP: Control-Plants only; CW+P: Control-Earthworms+Plants. LP: Pb-Plants only; LW+P: PbEarthworms+Plants. LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.7.4 Pb accumulation in plants
Sunflowers accumulated Pb in their tissues. Lead was much higher in plants from Pb
and Cs+Pb treatments than in controls. In roots, there was a significant difference in
Pb concentration between treatments (P=0.001) and in plants from ISMs with different
biota (P=0.004). Lead concentration was lower in plant roots in the Cs+Pb treatment
94
(874.83 ± 53.07 mg/kg) than in the Pb only treatment (1,187.05 ± 75.18 mg/kg). In
addition, Pb in plant roots was lower when earthworms+plants were in the ISMs
(913.63 ± 65.30 mg/kg) than plants only (1,148.24 ± 85.06 mg/kg) (Figure 3.35).
Lead concentration in roots and shoots
Lead concentration (ppm)
1800
1600
1400
1200
1000
Soil
800
Roots
600
Shoots
400
200
0
CP
CW+P
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.35 Lead concentrations in plants after 4 weeks
CP:
Control-Plants
only;
CW+P:
Control-Earthworms+Plants.
LP:
Pb-Plants
only;
LW+P:
Pb-
Earthworms+Plants. LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
Within Pb treatments, Pb in plant roots was higher in the ISMs with plants alone
(1,352.96 ± 56.43 mg/kg) than in those with earthworms+plants (1,021.13 ± 69.59
mg/kg). Within the Cs+Pb treatment, Pb in plant roots was also higher in when plants
alone were present than with earthworms+plants (943.53 ± 51.43, and 806.13 ± 85.75
mg/kg) (Figure 3.35).
Data for Pb concentrations in roots were transformed using log10. There was no
significant difference between treatments (P= 0.691). However, a significant effect of
biota was observed (P=0.022). In general, Pb in plant shoots was lower in ISMs with
earthworms+plants (22.14 ± 2.32 mg/kg) than in those with plants only (28.84 ± 3.33
mg/kg). When only plants were in the ISMs, Pb in shoots was higher in the Pb only
treatment (37.45 ± 4.80 mg/kg) compared with those from the Cs+Pb (20.22 ± 1.79
mg/kg) treatment. In contrast, when earthworms+plants were in ISMs, Pb in shoots
95
was lower in the Pb only treatment (15.38 ± 2.38 mg/kg) than in the Cs+Pb treatment
(28.90 ± 2.10 mg/kg) (Figure 3.35).
Lead concentration in roots (1,030.94 ± 60.01 mg/kg) was much higher than in shoots
(26.42 ± 3.85 mg/kg) (P<<0.001) (Figure 3.35).
3.3.7.5 Lead transfer factor in sunflowers
The results revealed that there was no significant difference in the Pb transfer factor in
sunflowers between treatments (P=0.880), and no effect of biota across treatments
(P=0.386). However, there was a significant difference within the Pb only (P=0.002)
and Cs+Pb (P=0.003) treatments. In the Pb only treatment, TF values were higher in
ISMs with plants only (0.03 ± 0.004) compared with those with earthworms+plants
(0.01 ± 0.002). Conversely, in the Cs+Pb treatment, TF values were higher in ISMs
with earthworms+plants (0.02 ± 0.001) than in those with plants alone (0.01 ± 0.001)
(Figure 3.36).
Lead transfer factor in sunflowers
0.035
Transfer factor
0.03
0.025
0.02
0.015
0.01
0.005
0
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.36 Lead transfer factor in plants after 4 weeks
LP:
Pb-Plants
only;
LW+P:
Pb-Earthworms+Plants.
LCP:
Cs+Pb-Plants
only;
LCW+P:
Cs+Pb-
Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
96
3.3.7.6 Lead translocation factor in sunflowers
Translocation factor
Lead translocation factor in sunflowers
0.045
0.04
0.035
0.03
0.025
0.02
0.015
0.01
0.005
0
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.37 Lead translocation factor in plants after 4 weeks
LP:
Pb-Plants
only;
LW+P:
Pb-Earthworms+Plants.
LCP:
Cs+Pb-Plants
only;
LCW+P:
Cs+Pb-
Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
There was significant difference in Pb translocation factor in plants between
treatments (P=0.049). Translocation factor was higher in the Cs+Pb treatment (0.03 ±
0.004) than in the Pb alone treatment (0.02 ± 0.004). However, there was no effect of
biota on translocation factor in plants across treatments (P=0.826). Within the Cs+Pb
treatment there was a significant difference (P=0.083) in the Pb translocation factor. In
addition, within the Pb only treatment translocation factor in ISMs with plants alone
(0.03 ± 0.003) was higher than in those with earthworms+plants (0.01 ± 0.004)
(P=0.034) (Figure 3.37).
3.3.7.7 Total uptake of Pb in sunflowers
There was no significant difference in total uptake of Pb in plants between treatments
(P=0.598) and no effect of biota across treatments (P=0.175). However, there was a
significant difference in total Pb uptake in plants within treatments (P=0.001 and
P=0.027 for Pb only and Cs+Pb treatments respectively). In the Pb single metal
treatment, total uptake of Pb was higher in plants from the ISMs with plants only (7.8
± 1.1 mg) than in plants from the ISMs with earthworms+plants (3.1 ± 0.4 mg).
97
Conversely, in the Cs+Pb treatment, the Pb taken up by plants in the earthworms+plant
ISMs was higher (5.8 ± 0.3 mg) than that taken up by plants from the plants only ISMs
(4.2 ± 0.5 mg) (Figure 3.38).
total uptake of lead in sunflowers
10
Total lead (mg)
8
6
4
2
0
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.38 Lead total uptake in sunflowers after 4 weeks
LP:
Pb-Plants
only;
LW+P:
Pb-Earthworms+Plants.
LCP:
Cs+Pb-Plants
only;
LCW+P:
Cs+Pb-
Earthworms+Plants.
(The bars are means of 4 replicates ± standard error)
3.3.7.8 Caesium accumulation by plants
Cesium concentration (mg/ kg)
Cesium concentration in sunflowers and soil
1400
Ceisum in top soil
1200
Roots
1000
Shoots
800
600
400
200
0
LCP
LCW+P
Biotas
Figure 3.39 Caesiumconcentrations inplants after 4 weeks
LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
98
Caesium was only measured in the Cs+Pb treatment. Data was transformed using 1/X
(where X was Cs concentration in plants) to achieve homogeneity of variances. The
analysis revealed the Cs level was higher in plants from ISMs with plants only (725.42
± 175.96 mg/kg) than in those from ISMs with earthworms+plants (604.93 ± 211.71
mg/kg) (P=0.031). Moreover, Cs was found to be higher in plant roots (1,041.61 ±
140.96 mg/kg) than in shoots (288.74 ± 26.16 mg/kg) (P<<0.05) (Figure 3.39).
3.3.7.9 Caesium transfer factor of sunflowers
The TF (transfer factor) of Cs from soil to sunflower shoots were not significantly
different between plants from ISMs with plants only and plants from ISMs with
earthworms+plants (P=0.127) (Figure 3.40).
Transfer factor (TF)
transfer factor for cesium in plants
1
0.9
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
LCP
LCW+P
Treatments
Figure 3.40Caesium transfer factor in plants after 4 weeks
LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.3.7.10 Caesium translocation factor
There was no significant difference in Cs translocation factor of plants between plants
from ISMs with plants alone and those with earthworms+plants (P=0.668) (Figure
3.41).
99
Cesium translocation factor in sunflowers
0.4
Translocation factor
0.35
0.3
0.25
0.2
0.15
0.1
0.05
0
LCP
LCW+P
Biotas
Figure 3.41Caesium translocation factor of plants after 4 weeks
LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.3.7.11 Total uptake of Cs in sunflowers
There was no significant difference between the total amount of Cs was taken up by
plants from ISMs with plants alone and those with earthworms+plants (P=0.450)
(Figure 3.42).
Total uptake of cesium by sunflowers
80
Total cesium (mg)
70
60
50
40
30
20
10
0
LCP
LCW+P
Biotas
Figure 3.42 Caesium total uptake in plants after 4 weeks
LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
100
3.3.7.12 Potassium in plants
Log10 was used to transform data for K in roots data. The data analysis revealed that
there was no significant difference between treatments (P=0.359) and no effect of
biota (P=0.182). Data for K in plant shoots was also transformed prior to analysis. The
ANOVA revealed that there was also no significant difference between treatments
(P=0.385) and no effect of biota (P=0.613). In addition, there was no significant
difference between K in roots and shoots (P=0.152) (Figure 3.43).
Potassium concentration (mg/ kg)
Potassium in plants
Roots
120000
Shoots
100000
80000
60000
40000
20000
0
CP
CW+P
LP
LW+P
LCP
LCW+P
Treatments
Figure 3.43 Potassium concentration in plants after 4 weeks
CP:
Control-Plants
only;
CW+P:
Control-Earthworms+Plants;
LP:
Pb-Plants
only;
LW+P:
Pb-
Earthworms+Plants; LCP: Cs+Pb-Plants only; LCW+P: Cs+Pb-Earthworms+Plants.
(The bars are means of 3 replicates ± standard error)
3.4 Estimation of the time required for effective phytoremediation
The equation from (Zhuang et al., 2005 cited in (Wuana and Okieimen, 2011)) was
adapted to estimate the duration of phytoremediation time tp(year). In the case for the
current study, the contaminated soil mass was 0.85 kg (0-5 cm top layer soils) and
sunflowers are assumed to be cropped 6 times. It was then assumed that the
contaminated soil by needed to be treated or phytoremediated to reduce the Cs
concentration from 250 mg/kg to 5 mg/kg and the Pb concentration from 1500 mg/kg
to 600 mg/kg for Pb. There are no guidelines for the concentration of Cs in soil
101
therefore 5 mg/kg of Cs was used since this is the “average” concentration in soil
(ATSDR, 2004). An alternative would have been using a set of background soil data to
estimate several percentile values and enable a comparison with actual background
levels. However this study was limited by time. Lead at 1500 mg/kg is classified as the
concentration in soils in industrial areas and 600 mg/kg is the concentration acceptable
for recreational open spaces, playgrounds, parks and secondary school (New South
Wales EPA, 1994, NEPM, 1999). The calculations revealed that phytoremediation is a
slow process which requires a long period of time. In addition, the presence of multicontaminants increased the time required for phytoremediation.
tp(year)
(Zhuang et al., 2005 cited in (Wuana and Okieimen, 2011))
Table 3.12 Estimation of time required for the phytoremediation of Cs and Pb using
sunflowers
Elements
Biota
Cs (years)
Without Pb
Pb (years)
With Pb
Without Cs
With Cs
Plants only
815±134
503±15
18,505±2,299
32,951±3,734
Earthworms+Plants
928±110
808±36
46,704±6002
22,401±1,410
102
Chapter 4. Discussion
The current study demonstrated that sunflowers and earthworms have the capacity to
extract Cs and Pb from contaminated soil as demonstrated by previous authors for
several species of plants including Nicotianatabacum L. (Mench et al., 1989, Kayser et
al., 2000),Brassica junceaL. (Huang and Cunningham, 1996)or Zea mays L. (Meers et
al., 2005)and several species of earthworms such as E. fetida, Lumbricusrubellus,
Lumbricuscastaneus,
Lumbricusterrestris,
for
Pb(Nahmani
et
al.,
2007)and
Octolasiumlacteumfor Cs (Brown and Bell, 1995, Spurgeon et al., 1994). However the
capacity was dependent on a number of abiotic and biotic characteristics as have been
presented in the results chapter for which the explanations are discussed here.
4.1 Integrated soil microcosms
In the current project Integrated Soil Microcosms (ISMs) were created to evaluate the
fate and accumulation of Cs and Pb in soil combined with biota. ISMs represent the
natural ecosystem in that they contain soil and biota and in this study, successfully
provided information on how each component of the abiotic and biotic components
were affected by the presence of the trace metal contaminant and how each component
was in turn affected by and altered the response to the metals. Such ISMs have been
used previously to assess the toxicity of pesticides (Burrows and Edwards, 2002) and
the effects of trace metals (Edwards, 2002) with similar outcomes to the current study.
The use of these ISMs enabled the evaluation of many factors which included the
effects of Cs and Pb on soil properties (pH, conductivity, moisture, and organic matter)
and their mobility in soil. Earthworm mortality, distribution, biomass and
accumulation, plant germination, growth, biomass and accumulation were all
evaluated within a representative mini ecosystem. The leaching of Cs and Pb with time
was also incorporated. In addition it was possible to incorporate a mixture of Pb and
Cs in the ISMs to simulate the situation in nature where most contaminated soils
contain more than one metal and synergistic effects between pollutants may occur
(Nahmani et al., 2007). Testing a single species in isolation does not reflect the natural
103
effects of toxicants in an ecosystem and the ISMs could be considered more
representative.
4.2 Change of soil properties and metals in soil during the experiments
The soil selected for the study was very suitable for use in microcosms based on the
results of the soil analysis (Table 3.1). The pH=7.2 of initial soil was neutral and
suitable for both earthworms (E. andrei) as reviewed by (Widmer, 2010) and for the
optimal growth of sunflowers (Putnam et al., 1990). Earthworms can be found in
acidic soils of pH 4 but are more abundant when pH is close to neutral (pH 7) (Wild,
1993b). The soil moisture (25%) and organic matter (11%) was suitable for plants and
earthworms to survive. The soil type was found to be a clay loam which is suitable for
sunflowers (from sandy loam to clay). According to Giannakopoulou et al. (2007),
whatever the pH values, higher Cs sorption was observed in the clay soil, followed by
the clay loam and the loam soil. The minimum sorption was found to be in sandy loam
soil (Giannakopoulou et al., 2007). In general, this soil was considered suitable for the
study to determine bioremediation of trace metals, compared with OECD standard soil
which has 70% sand, 20% clay, and 10% organic matter (OECD, 1984); since the
latter would not retain trace metals to the same extent.
When Cs was added to the soil (at 25 mg/kg and 250 mg/kg), there was a minor
increase in soil pH above that of the initial soil (pH=7.2). This may be due to the fact
that plants take up excess anions from soils thus raising the soil pH (Wild, 1993). In
addition, the activity of earthworms can increase soil pH by the production of excreta.
However, all pH values were alkaline and would therefore have no effect on Cs
accumulation by plants (Wild, 1993).Ebbs et al., 1998 suggested that a soil pH<5.5
would be required to convert Uranium to its most phyto-available form in soil. Wild
(1993b), (Monna et al., 2009) and (Giannakopoulou et al., 2007) proved that when the
pH increases the adsorption of heavy metals and essential nutrients into soil
compartments increased. Trace metals can be precipitated, adsorbed by the negatively
charged particles, held on surfaces of clay, hydrated oxides and humus. The maximum
Cs sorption in soils was observed at soil pH=8 (Giannakopoupou et al., 2007). The
desorption of trace metals and their concentration in solution are highest in acid soils
104
(Wild, 1993), hence in the current study, since the soils were alkaline it is not
surprising that the bioavailability of the two trace metals was low.
With the Pb alone and Cs+Pb contaminated soils, soil pH was also increased compared
to initial soil with the lowest pH being 6.4 ± 0.05. Again Wild stated that liming acidic
soils to raise pH to 6.5 reduces the availability of most metals to plants (noting that it
not necessary to express soil pH measurements more accurately than to the nearest 0.2
division of a unit Marginal soil alkalinity may be a reason for the limited uptake of
trace metals by sunflowers in the current study.
Soil moisture was 26-40%; 25-36% in the top and 32-43%; 30-40% at the bottom for
the Cs and Cs+Pb experiment respectively (Table 3.4 and 3.9). These are suitable
conditions for earthworms to remain active. Olson (1928) found that the greatest
number of earthworms occurred in soil containing moisture from 12% to 35% (Grant,
1955).
Soil organic matter remained approximately 11-12%. This was a relatively high
organic matter content. The organic matter content was expected to be increased by the
activities of the earthworms (Hendrix, et al., 1984 and Scheu and Parkinson, 1994)
however the experiments were only run for four weeks and within this short period no
increase in the soil organic matter content was evident.
4.2.1 Caesium in soils:
Most of Cs was remained in the top layer soils. The method of spiking the soil was
adequate to ensure this in both experiments. This may due to the high affinity of Cs
with organic matter and clay particles in soils. However, the presence of Cs in the
bottom layer soils after 4 weeks proved that Cs had moved through the soil profile.
The higher Cs concentration used for soil contamination, the higher the level of Cs in
the bottom layer soils, demonstrating that the mobility of Cs through the soil was
directly related to the concentration of the metal in the top soil (Figure 3.1).
105
The presence of earthworms increased the Cs concentration of the bottom layer soils in
both the 25 mg/kg and 250 mg/kg Cs treatments (Figure 3.1). This may be a result of
earthworm casts and movements mobilizing the Cs. In addition, earthworms create
tunnels through the soil core therefore Cs may move down with water during
irrigation. The spiked solution might also move down at the beginning as a result of
soil porosity. This result was different to that of Knox et al., 2001 who reported that
the vertical migration of Sr, Cs, Pu, and Am is expected to be 0.3 to 0.5 cm per year
(Knox et al., 2001). It is probable that the soil used in the current study was more
porous and had a lower clay and organic matter content than the soils studied by Knox
et al (2001). This may also be due to the added Cs being quite mobile unlike Cs that
already exists in soil naturally.
After four weeks, the concentration of Cs in the soil did not change much. This is not
an extensive time for remediation and the obvious results were not expected. In
addition, the presence of earthworms did not help to reduce the Cs in soils (Figure
3.1). Whether Cs in soil changes over longer periods of time in similar experiments
remain to be determined. The explanations might come from soil pH. When soil pH >
7.0 the Cs sorption is increased because of the creation of variable charges on clay
minerals and organic matter (Monna et al., 2009). As mentioned above the maximum
sorption of Cs was at soil pH=8. Hence in the current study, with the soil pH being
high, Cs in soils altered only marginally.
The potassium concentration in soil was not different when soil was contaminated with
Cs. Cs is known to compete with potassium for uptake by plants. Bunzl et al., 2000
revealed that there was a negative relationship between 40K in the soil and 137Cs in the
plants (Bunzl et al., 2000). However, similar competition was not evident in the results
of the current study (Figure 3.17 and 3.18), possibly because the experimental period
was short.
106
4.2.2 Lead in soils:
The presence of Cs did not affect the Pb removed from contaminated soils after 4
weeks since the concentration remained the same when compared with that of the
initial soil (Figure 3.16). The presence of earthworms did not increase the
phytoremediation of Pb in soils. Moreover, most of Pb remained in the top layer of soil
after four weeks (Figure3.16). Lead is known to be very stable in soil. It attaches
tightly to the soil particles especially clay and organic matter (Lounamaa, 1956,
Zimdahl and Skogerboe, 1977, as cited in (Alloway, 1995, Sipos et al., 2005). The soil
used in this experiment has very high percentage of clay (36%) which is an
explanation for the Pb remaining in the top soil layer. The other factor that can affect
Pb removability is soil pH. In this experiment soil pH decreased to around 6.5 (Table
3.7) but did not become more acidic, therefore the bioavailability of Pb did not
increase during the four week experiment. Metal cations are more mobile under acidic
condition (pH<5.5) and especial with Pb it becomes active at pH 4 (GWRTAC, 1997)
(Herms and Brumner, 1984, as cited in (Karaca et al., 2010), but as discussed the soil
pH in the current
study remained high thus not enhancing the mobility and
bioavailability of Pb.
Interestingly however, Pb moved through to the bottom layer soil and was detected at
low concentrations in this layer (Figure 3.16). This indicates some mobility of Pb
following irrigation of soils. It was noticeable that in the Treatment with Pb alone with
the presence of earthworms only (183 mg/kg, approximately 12%), the concentration
of Pb in bottom layer soils was higher than when there was nobiota present (90 mg/kg,
about 5%), plants only (134 mg/kg, ~9%) and earthworms+plants (130 mg/kg, ~9%).
The earthworms clearly enhanced the Pb in the bottom layer by their tunnel systems,
casts, and movement of earthworms up and down soil core as reported by (Bouché,
1972, 1977 cited in (Brown and Bell, 1995). The authors explained that by increasing
the soil porosity and vertical burrows, earthworms would increase the infiltration rate
and deep penetration of contaminants on the surface soil. This can be used to explain
the presence of Pb in the bottom layer soil. Furthermore, through the surface-feeding
activities, earthworms will mix and deepen the contaminated surface layer.
It is also worth noting that the lead accumulation in this experiment is from soluble Pb
added to soil that maybe different from naturally occurring forms of Pb. Thus the
107
experimental results could differ from what would be seen in naturally contaminated
soils.
4.3 Leachates
In the Cs experiment, all leachate pH values were ≥ 7. It is clear that the earthworm
activities altered the pH of leachates. Cs was detected in the leachates from the higher
concentration treatment of 250 mg/kg Cs (Figure 3.2). This indicates that Cs can move
through the soil profile and may contaminate the groundwater. However, the soil pH in
the current study was higher than 7 therefore the availability of Cs was limited. Hence,
there was no toxicity detected when the leachates were tested using the microtoxtest.
In the Cs+Pb experiment Cs was also measured at a very low concentration in
leachates and the presence of Pb did not affect Cs in leachate (Figure 3.20).
In addition, Pb was also detectable in leachates after the short time of four weeks
(Figure 3.19). With Pb, the concentration in leachates was higher in the treatment with
Pb alone compared with the Cs+Pb treatment. It is possible that Cs interfered with the
mobility of Pb, since it seemed to be generally more mobile than Pb in the
experimental soil. It was apparent that Cs concentration was higher than Pb in
leachates (about 0.2 mg/kg and 0.1 mg/kg respectively). In the no biota treatment, the
leachates had the lowest concentration of Pb. This may be due to the lack of
earthworm activities and the root systems of the plants in enhancing the leaching of
Pb. Thus the added metals were more mobile and available as reported by (Spurgeon
and Hopkin, 1995, Davies et al., 2003) in similar studies with Pb contaminated soils.
4.4 Earthworms
4.4.1 Mortality and mobility of earthworms
In the Cs experiment, there were more earthworms dead in 250 mg/kg treatment than
in the 25 mg/kg treatment or the control (Figure 3.3). It is possible that the higher
concentration of Cs was toxic to earthworms. The distribution of earthworms was not
affected by Cs (Figure 3.4). The number of earthworms in top and bottom layer soils
was not significantly difference. More interestingly, earthworms in most cases gained
weight in exposures with Cs except in the 250 mg/kg treatment with
earthworms+plants (Figure 3.5). They lost weight in this exposure (approximately 4%
108
and 16% for 2 and 4 weeks respectively). This result was not in agreement with the
study of Domínguez-Crespo et al. (2012). In their study, there was a decrease in
earthworm biomass which was affected by the quantity of food (Domínguez-Crespo et
al., 2012) . However, in the current study, it would seem that the earthworms preferred
the Cs contaminated soils and ingested the soil.
In the Cs+Pb experiment, earthworm mortality was not affected by the presence of Pb
only or Cs+Pb (Figure 3.22). Earthworms have a high tolerance to heavy metals. This
was proved by Spurgeon et al., 1994 who demonstrated that earthworms can survive
and reproduce in anthropogenically contaminated soils with high concentrations of
trace metals. In his experiment with Cd, Zn, and Pb, Pb was the metal of the least
toxicityto E. fetida. Significant mortality was recorded only at 10,000 mg/kg Pb. The
LC50 of 14 and 56 days was 4,480 mg/kg and 3,760 mg/kg respectively (Spurgeon et
al., 1994). In the current study the distribution of earthworms was not influenced by
the chemicals (Figure 3.23). They did not avoid the top soil which was contaminated
in the current study and were found to be distributed equally between top and bottom
layer soils. Moreover, they seem prefer the presence of Cs+Pb since higher numbers of
earthworms was found in the top layer soils, indicating that the concentrations of
metals used in the Cs+Pb experiment were not unpalatable to worms
Earthworm biomass decreased in all treatments (Figure 3.24). The reasons were the
lack of food for earthworms during 4 weeks. Earthworms were fed in nutrient rich
conditions before the experiment. This result was similar to that of Spurgeon et al.,
1994. The authors indicated that the weight of earthworms declined in all treatments
(including controls) during the experiment and the explanation was the lack of suitable
food in OECD standard soil medium used (containing up to 10% of organic matter)
and suggested that animal manure be added to the soil. The soil used in the current
study had a similar % of organic matter (i.e. 11%) but the clay content was much
higher than that of the OECD standard soil, and no manure was added, leading to the
lack of food.
Overall, the mortality and reduction in the biomass of earthworms can be explained by
the lack of food during the study. On the other hand, introduction of high populations
109
of earthworm into soil cores can also cause intense earthworm activity and result in
death or loss weight (Edwards, 2004).
4.4.2 Caesium and Pb accumulation by earthworms
Earthworms accumulated Cs and Pb in substantial amounts. The concentration of Cs in
earthworms increased with increase in Cs in contaminated soil (Figure 3.6). There was
a positive correlation in Cs concentration in soil and earthworms. When the Cs
concentration in soil increased 10 times (25 mg/kg to 250 mg/kg Cs) Cs concentrations
in earthworms after 4 weeks increased 10 to 11 times regardless of whether the worms
were sampled from top or bottom layer soils (Figure 3.6).
With change in time from 2 to 4 weeks, the concentration of Cs in earthworms
declined nearly twice in the 25 mg/kg Cs treatment but not in the 250 mg/kg Cs
treatment (Figure 3.6). This maybe due to the fact that the low amount of Cs
bioaccumulated in the first 2 weeks in the 25 mg/kg treatment was more readily
excreted. The greater amount of Cs bioaccumulated in the 250 mg/kg Cs would have
been distributed into less labile pools within the earthworm. The accumulation of Cs
was higher in the earthworms in the top soil layer in the 25 mg/kg Cs treatment but not
in the 250 mg/kg treatment (Figure 3.6). However the TF from soil to earthworms was
apparently higher in the bottom soil (Fig 3.7). This is probably a result of earthworms
naturally feeding in the top soil layer but moving to the bottom layer to avoid the
contaminated topsoil and remaining there when microcosms were sampled. The
apparent similar uptake of Cs by earthworms from top and bottom soils is also
probably due to earthworm movements. This was similar for Pb (Fig 3.26 and Fig
3.27). Earthworms appear to accumulate much higher levels of metals from polluted
soils than most other soil animals (Beyer et al., 1982). This was confirmed in the
current study. The transfer factor (TF) can also be used to indicate the accumulation
ability of earthworms. In this study, the highest TF value was observed to be nearly 1.0
of top layer earthworms. Hence, the highest amount of Cs removed was approximately
10.3% (25.8 mg) in the Cs experiment from the 250mg/kg treatment worms (Figure
3.7 and 3.8). In contrast, in the Cs+Pb experiment, the Cs accumulation was reduced.
The highest TF values of earthworms were 0.3 in the presence of earthworms+plants
110
(Figure 3.29) and the highest total Cs uptake was 4.2 mg/kg equal to 1.7% of Cs in soil
(Figure 3.30). This may due to the presence of Pb in the mixture experiment.
The concentration of potassium was measured in earthworms since previous
researchers (Hedrich and Dietrich (1996); Zhu and Smolders (2000) cited in Soudek et
al., 2004) have noted changes in K concentrations when biotas are exposed to Cs. K in
earthworms was not altered in the microcosms treated with Pb alone, though it was
reduced in the Cs+Pb treatments (Figure 3.31). It was obviously that the potassium
concentration was lower in Cs+Pb exposed worms compared with controls and the
treatment with Pb alone. Therefore in the current study the competition between Cs
and K noted by some previous researchers (Hedrich and Dietrich (1996); Zhu and
Smolders (2000) cited in Soudek et al., 2004) was evident. However the lowering of K
when Cs and Pb were present in combination could also be explained by the
competition between Pb and K for uptake by earthworms or by competition between
Pb and Cs having a secondary effect on K uptake. It is also possible that Pb being a
non-essential toxic metal caused a toxic effect on some metabolic pathways for K
including nerve function.
Pb accumulation in earthworms was reduced by the presence of Cs in soils (Figure
3.25). The TF values were significantly lower in Cs+Pb earthworms (0.2) compared
with earthworms in treatments with Pb alone (0.3) (Figure 3.26). This was followed
with a lower Pb total uptake of 12.4 mg compared with 13.4 mg (Figure 3.27). In
addition, Pb concentrations were higher in worms from the top soil layer (Figure 3.25).
Overall, the presence of Cs+Pb as a mixture had adverse effects on Cs and Pb
accumulation by earthworms, thus reducing bioremediation of the trace metals from
soils.
4.5 Plants
4.5.1 Plant germination, growth and biomass.
Sunflowers could germinate well in the experimental soil, even after contamination
with Cs and Pb, and survived for the entire experimental period (Figure 3.9). There
111
were no symptoms of diseases and nutrient deficiency. All plants looked healthy
throughout the experiment. This proved that sunflowers have a high tolerance with Pb
and Cs as reported by (Tang et al., 2003, Schmidt, 2003b, Pilon-Smits, 2005). The
presence of Pb only and Cs+Pb did not affect the germination of the sunflowers
(Figure 3.32). This may be because of the permeability of sunflower seed coats and the
buffering capacity of soil. Soil buffering can reduce the potential bioavailability of
contaminants in the environment which in turn reduces the toxicity (Hatzinger and
Alexander, 1995, Vig et al., 2002).
The growth of plants in most of the contaminated soils seemed slightly reduced
compared with the controls. However, the effect was not significant (Figure 3.10 and
3.33). This can be explained by the short experimental period and the number of
replicate too limited to detect a significant difference. Pb absorption by the roots could
result in reducing the growth of plant (Godbold and Kettner, 1991). These authors
explained that the emergence and growth of secondary roots are especially sensitive to
Pb toxicity (Godbold and Kettner, 1991). Pb can block the uptake of K, Ca, Mg, Mn,
Zn, Cu, and Fe by modifying membranes activity and permeability and making them
unavailable for uptake and transfer to plant (Patra et al., 2004). High concentration of
Pb will result in water deficit and transpiration rate reduction, cell sap osmotic
pressure and xylem water potential alteration and causing negative effect on water
situation of plants (Parys et al., 1998). Pb can inhibit enzymes and protein
conformation by binding to -SH and -COOH groups (Sharma and Dubey, 2005). In
addition, Pb can replace some enzyme cofactors such as Mn and Mg, therefore
interfering with photosynthesis and electron transportation; by affecting oxygen
evolution and chlorophyll levels and causing alteration in thylakoid membranes
structure (Patra et al. 2004). Pb also inhibits the key chlorophyll biosynthesis enzymes
and Calvin cycle enzymes (RuBisCO, phosphoenal pyruvate carboxylase, and ribulose
5-phosphate kinase) thereby reducing the CO2 fixation rate and efficiency (Sharma and
Dubey 2005). One special effect of Pb is cell cycle disruption by interfering with the
alignment of microtubules on the mitotic spindle (Eun et al., 2000). If the current study
had run the Pb exposure for a longer period than four weeks such results may have
been evident.
112
The presence of Pb and Cs+Pb in a mixture did not affect plant dried biomass (Figure
3.34). However, the higher concentration of Cs induced an increase in the plant
biomass (Figure 3.11). This may be due to the similarity of the Cs uptake pathway
with that of the essential nutrient potassium, therefore a higher concentration of Cs
was taken up and also stimulated plant growth. Sunflowers are considered a high
biomass producing species (Kumar et al., 1995). Therefore, they were considered to be
suitable for bioremediation of metals since a large biomass is very important to ensure
maximum uptake of metals from contaminated soils. In this experiment earthworms
were introduced with the expectation that they can increase plant growth and biomass
by increasing soil structure, decreasing soil bulk density and breaking down
undecomposed litter as described by(Noble et al., 1970, Lavelle, 1997). However, the
presence of earthworms did not show any benefit in improving plant growth and
biomass. This could be a result of the short time of the experiment. Four weeks is not
long enough to ensure that earthworms turn over undecomposed organic matter in soil
into decomposed organic matter and then induce the plant growth and biomass as
predicted by Edwards, 2004.
Radiocaesium remains mostly in top (0-8 cm) soils (Mikhaylovskaya et al., 1993),
therefore the root length of plants plays an important role in phytoremediation. In the
current study, the root length of the sunflower (H. annus) plants was between 4 and 9
cm and therefore may be considered suitable for radiocaesium uptake and
decontamination of sites contaminated with radiocaesium.
4.5.2 Caesium and Pb accumulation by plants:
Considering Cs, the higher the concentration in soil the higher the concentration
detected in shoots and roots (Figure3.12). This result was opposite to that of Bunzl et
al. (2000) which indicated that increased 137Cs content in plant was not associated with
increased
137
Cs content in soil (Bunzl et al., 2000). Conversely, Tang et al (2011)
stated that shoot and root concentration of Cs were significantly increased with
increasing concentration of Cs in soil (Tang et al., 2011). In the study of Tang et al., it
was found that in Amaranthuscruentus L. shoot Cs concentration ranged from 11,000
113
to 17,000 mg/kg in soil treated with 1000 mg/kg Cs, and from 7000 to 12,000 mg/kg
in Phytolaccaamericana Linn. exposed to the same Cs concentration. In roots the
values were 9500 to 13,000 mg/kg for A. cruentus L. and 7500 to 11,000 mg/kg for P.
americana Linn. TFs values ranged from 7.65 to 20.34 for P.americanaLinn. and
11.34 to 42.31 for A.cruentus L. .Concentration ratios (translocation factor) were 0.63
to 1.1 and 1.17 to 1.41 for P. Americana Linn. and A. cruentus L. ,respectively. In the
current study, high TF values of 1.9and 4.0 for 25 mg/kg and 250 mg/kg treatments
respectively were observed after 2 weeks (Figure 3.13). The difference in TF values
from that of the study by Bunzl et al (2000) can be explained by the possibility that the
transfer factor of stable Cs may be lower than that of 137Cs as suggested by (Borghei et
al., 2011). In the current study the highest Cs concentration in shoots was 1138 mg/kg
and in roots 3439 mg/kg for plants in the 250 mg/kg treatment. This result was very
close to that of Tang et al., 2011 demonstrating that sunflower (H. annus) is as
effective as A.cruentus L. and P.americana.in phytoremediation of Cs. The
translocation of Cs from roots to shoots was very high (approximately 75%) (Figure
3.14). Furthermore, Cs total uptake was up to 98mg in earthworms+plants from the
250 mg/kg treatment at the 4 week sampling (Figure 3.15). The transfer of Cs from
soil to plant was also influenced by plant characteristics. Broadley and Willey (1997)
observed up to 100 fold differences in Cs accumulation of 30 plant taxa (Broadley and
Willey, 1997). In the current study the presence of Pb did not influence uptake of Cs.
This result was similar to that of Scottiet al., 2000 who stated that the uptake of
strontium and Cs was not affected by the presence of heavy metals in either leaves or
in fruits, or their translocation from leaves to edible parts (Scotti and Carini, 2000).
However, the TF was reduced to 0.9 for 250 mg/kg Cs in the presence of Pb (Figure
3.40).
There was a significant decrease of Cs in roots of 250 mg/kg treatment (Figure 3.12).
The reducing of Cs accumulation with time is probably due to “ageing effect”, in
which more Cs was locked into the soil (interlayer spaces of clay mineral) and organic
matter with time and was less available for root accumulation (Cook et al., 2007,
Varskog et al., 1994, Willey and Martin, 1995, Tsukada et al., 2002).
114
Lead accumulation was found to be higher in roots than in shoots in plants grown in
Pb contaminated soils. The Pb concentration was 1,030.94 ± 60.01 mg/kg in roots and
26.42 ± 3.85 mg/kg in shoots, respectively when plants were grown in soil
contaminated with 1500mg/kg of Pb (Figure 3.35). This result was similar to that of
many previous studies. For example, Jusselme found that Pb concentration was at least
5 times higher in roots than in shoots of Lantana camara (Jusselme et al., 2012).
In the current study, Pb was lower in plants in the presence of earthworms from both
Pb alone and Cs+Pb treatments. In roots the Pb concentration was 913.63 ± 65.30
mg/kg and 1,148.24 ± 85.06 mg/kg and in shoots was 22.14 ± 2.32 mg/kg and 28.84 ±
3.33 mg/kg respectively for plants from earthworms+plants and plants only treatments.
The addition of earthworms was expected to enhance the bioavailability of metals for
plant uptake, however the result was in contrast to that of Jusselme et al., (2012)
(Figure 3.35). Their study on Pb accumulation from soil by L.camarain the presence of
the earthworm Pontoscolexcorethrurus indicated that the presence of earthworms
enhanced the weight of plant shoots and roots and Pb uptake by the plants. Total Pb
uptake in the shoots and roots was 3 and 2 times higher in 500 mg/kg and 1000 mg/kg
Pb in soilin the presence of earthworms than in their absence (Jusselme et al., 2012).
It has been suggested that a reason for the decrease in metal bioavailability of wormworked soil is that earthworms directly take up water-soluble metals in soil solutions.
This could directly affect the amount of metals in soil for plants and therefore reduce
metal concentrations in soil solution. This reduction may cause an increase in the pHdependent surface charge density and result in more metals binding to colloids. As a
result, the metal available for plants to take up decreased (Cao et al., 2001, Shan et al.,
2002, Wen et al., 2004a). Conversely, reasons for increases in observed metal
availability are also suggested to be due to the earthworm activities (release of metal
chelating organic materials forming organo-metal complexes) and decomposition of
earthworm tissues themselves with high metal burdens making these metals more
available than they were in the soil(Ireland, 1975, Ma et al., 2002, Currie et al., 2005,
Wang et al., 2006).
115
The presence of Cs reduced Pb accumulation in plants. The Pb concentration in plants
was 874.83 ± 53.07 mg/kg and 1,187.05 ± 75.18 mg/kg for Cs+Pb and Pb alone
treatments respectively (Figure 3.35). The TF values were 0.02 and 0.03 for with and
without Cs (Figure 3.36). These TF values were far below that of the 1.7 reported for
Brassica juncea (Indian Mustard)found in 500 mg/kg Pb contaminated soils (USEPA,
2000). However, sunflower is considered a high biomass crops and therefore can be
used forphytoremediation of contaminated soils.
4.6 Remediation efficiency
Total Cs and Pb remediated was calculated based on the highest amount of Cs and Pb
taken up by earthworms, and above ground shoots of the sunflower plants in the table
below;
Table 4.1 Cs and Pb accumulation efficiency
Cs
Without Pb
Earthworms
Biota
Plants
Total uptake in 4 weeks
Time
estimation
(years)
Plants only
Earthworms+Plants
25.76 mg
(10%)
97.68 mg
(39%)
123.44 mg
(49%)
815±134
928±110
Pb
Including
Pb
4.16 mg
(2%)
61.82 mg
(25%)
65.98 mg
(27%)
503±15
808±36
Without Cs
Including Cs
13.41 mg
(5%)
7.79 mg
(0.5%)
21.20 mg
(5.5%)
18,505±2,299
12.41 mg
(0.5%)
5.84 mg
(0.4%)
18.25 mg
(0.9%)
32,951±3,734
46,704±6002
22,401±1,410
Comparing the Cs and Pb concentration and transfer factor in earthworms and plants;
Cs and Pb in leachates, and in bottom layer soils, it could be interpreted that Cs was
more mobile and bioavailable than Pb. In addition, based on a time estimation, it
would take longer to remediate Pb than Cs in contaminated soils as calculated based
on the results on the current study in Table 3.12. Moreover, the remediation process
for Pb would take even longer in the presence of Cs and possibly other metals as
multi-contaminants. It was obvious that Cs was much easier to phytoremediate than Pb
using sunflowers. Using earthworms as a bio-amendment was a reasonable approach
in terms of environmental sustainability though in the current study the enhancement
of phytoremediation in the presence of earthworms was not evident. The use of
integrated soil microcosms with the plants and earthworms and the collection of
116
leachates were useful in evaluating the fate and accumulation of Cs and Pb in
contaminated soil as representative mini ecosystems in the current study.
117
5. Conclusions
Integrated soil microcosms were demonstrated to be an effective tool in addressing the
fate and accumulation of Cs and Pb using a combination of earthworms and plants in
this study.
Soil pH increased during 4 week experimental period. However, soil organic matter
was not altered at the end of experiment. Soil moisture was suitable for plant growth
and earthworm activity and was maintained throughout the 4 week experimental
period. Cs and Pb moved through the soil core in 4 weeks with Cs being more mobile
than Pb. Cs and Pb was leached to leachates with pH>7 after 4 weeks and shown no
toxicity on Vibrio fischeri.
Earthworms accumulated both Cs and Pb in their body tissues. When Cs and Pb were
present as a mixture each reduced the accumulation of the other metal by earthworms.
The concentration of Cs and Pb was higher in the earthworms in the top soil layer. Cs
and Pb concentrations accumulated by earthworms were directly related to the
concentration of metals in the soil. However, Cs concentrations in earthworms reduced
from the 2 week sampling to the 4 week sampling. Earthworms tolerated the
experimental concentrations (25 and 250 mg/kg and 1500 mg/kg) of Cs and Pb
respectively, without avoidance reactions.
Plant could geminate, survive and growth under the experimental concentrations of Cs
and Pb. The presence of earthworms did not have any positive effects on plant
parameters. Plants accumulated Cs and Pb with higher TF of Cs than Pb. The
concentration of Cs and Pb was higher in plant roots than in shoots. At the higher
concentration of Cs in soil, higher concentrations of Cs in plants were observed.
However, there was no difference in the Cs accumulation of sunflowers at 2 and 4
weeks. In addition, when Pb and Cs were in the soil as a mixture, the accumulation of
both Pb and Cs were reduced. The presence of earthworms did not enhance the
accumulation of Cs and Pb.
118
Using integrated soil microcosms to investigate the fate and accumulation of Cs and
Pb especially and trace metals in general together with biota amendments (such as
earthworms) is a good approach and needs further investigation. There is inadequate
information on both Cs and Pb in the terrestrial environment, especially as a mixture.
This study has demonstrated that Cs and Pb interfere with each other and compete for
uptake by earthworms and plants, thus reducing overall bioremediation. Therefore,
such multi-element exposure experiments are required to understand competition
between trace metals; in order achieve effective bioremediation of contaminated soils.
119
6. Future work
In the current study the presence of earthworms did not have a positive effect on plant
parameters including Pb and Cs accumulation. However, it is possible that reasons for
this enhancement not being evident (i.e. not statistically significant) was due to the
limited number of replicates in the complex experimental set up; and because the
length of the experiment was only four weeks. It would therefore be beneficial in
future work if the number of replicates and sampling occasions were increased as well
as the experimental period extended.
In addition, in the current study it was not possible due to time limitations to measure
one of the important biotic components that should be measured i.e. microorganism
activity. Microorganism activities could be impacted by toxicants in single or mixture
forms and thus affects bioremediation. Therefore, using microorganism enzyme
activities to evaluate the effects of toxicants and whether this in turn affects
bioremediation, will be of benefit.
The sunflower is a good candidate for phytoremediation because of its high tolerance
and high biomass production (Tang et al., 2003, Schmidt, 2003, Pilon-Smits, 2005).
Therefore, using sunflowers in combination with the addition of different species of
earthworms or humic substance and soil fungi or micoorganisms to evaluate if these
additional factors could improve the phytoremediation ability of
sunflowers is
neccessary.
The laboratory microcosm experiments needs to be extended to further studies in field
conditions so that the data are more realistic and applicable, since field conditions are
often very different to those in the laboratory. The synthetic solutions of trace metals
used in laboratory experiments could be more bioavailable than those in the field
found in actual contaminated soils.
120
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