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BIOGEOCHEMICAL BEHAVIOUR OF HEAVY METALS IN SOIL-PLANT SYSTEM
By
Muhammad Shahid
Department of Environmental Sciences COMSATS Institute of Information Technology, Vehari
Pakistan
HIGHER EDUCATION COMMISSION
ISLAMABAD – PAKISTAN
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institutions in the country. For information please visit: www.hec.gov.pk
HEC – Cataloging in Publication (CIP Data):
Muhammad Shahid, Dr.
Biogeochemical behavior of heavy metals in soil-plant system
1. Soil and Environmental Science
577.14 – dc23 2017
ISBN: 978-969-417-195-1
First Edition: 2017
Copies Printed: 500
Published By: Higher Education Commission - Pakistan
Disclaimer: The publisher has used its best efforts for this publication through a rigorous system of
evaluation and quality standards, but does not assume, and hereby disclaims, any liability to any
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Copyrights @ Higher Education Commission
Islamabad
Lahore Karachi Peshawar Quetta
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DEDICATION
To ALLAH Almighty
Holy Prophet Muhammad (P.B.U.H)
My Parents & Teachers
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TABLE OF CONTENTS
LIST OF TABLES …………………………………………………………………………………………xi
LIST OF FIGURES ………………………………………………………………………………………xiii
LIST OF ABBREVIATIONS ……………………………………………………………………….…….xv
PREFACE ………………………………………………………………………………………………..xvii
FOREWORD ………………………………………………………………………………………….... xix
ACKNOWLEDGEMENT ……………………………………………………………………………......xxi
POLLUTION OF HEAVY METALS AND METALLOIDS............................................................. 1
1.1 METALS AND TYPES …………………………………………………………………………….... 2
1.2 HEAVY METALS AND METALLOIDS ……………………………………………………………. 4
1.3 HEAVY METAL SOIL CONTAMINATION ………………………………………………………… 5
1.4 HEAVY METALS OCCURENCE IN SOILS ………………………………………………………. 6
1.5 POTENTIAL HAZARDS OF HEAVY METAL CONTAMINATION ……………………………… 8
1.5.1 ENVIRONMENTAL PERSISTENCE …………………………………………………………. 8
1.5.2 POTENTIAL PHYTO-TOXICITY ……………………………………………………………… 8
1.5.3 POTENTIAL RISK OF METALS TO HUMAN HEALTH ……………………………………. 9
BIOGEOCHEMICAL BEHAVIOUR OF ARSENIC IN SOIL-PLANT SYSTEM …………………. 11
2.1 INTRODUCTION …………………………………………………………………………………… 11
2.2 GLOBAL USES OF ARSENIC ………………………………………………………………….... 13
2.3 ARSENIC LEVELS AND SOURCES IN SOILS ………………………………………………… 14
2.4 PHYTOAVAILABILITY OF ARSENIC IN SOILS ……………………………………………….. 15
2.5 EFFECT OF SOIL CHEMICAL PROPERTIES ON ARSENIC PHYTOAVAILABILITY .…… 15
2.5.1 SOIL PH AND ARSENIC PHYTOAVAILABILITY ……………………………………………. 15
2.5.2 SOIL ORGANIC MATTER AND ARSENIC PHYTOAVAILABILITY .………………………. 16
2.5.3 SOIL MICROBES AND ARSENIC PHYTOAVAILABILITY .………………………………… 17
2.6 SOIL-PLANT TRANSFER OF ARSENIC ………………………………………………………. 18
2.6.1 MOLECULAR UNDERSTANDING OF ARSENIC ABSORPTION BY PLANTS ……… 18
2.6.2 ARSENIC AND COMPLEMENTARY CATIONS .………………………………………….. 18
2.6.3 ARSENIC SEQUESTRATION INTO PLANT ROOTS ……………………………………. 19
2.6.4 ARSENIC TRANSLOCATION TO PLANT SHOOTS .…………………………………….. 19
2.6.5 ARSENIC ACCUMULATION IN EDIBLE PLANT PARTS ……………………………….. 20
2.7 TOXIC EFFECTS OF ARSENIC ON PLANTS …………………………………………….... 20
2.7.1 ARSENIC TOXICITY TO PLANT GROWTH ………………………………………………. 21
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2.7.2 ARSENIC GENOTOXICITY …………………………………………………………………. 21
2.7.3 ARSENIC-INDUCED OXIDATIVE STRESS ………………………………………………. 22
2.8 ARSENIC DETOXIFICATION IN PLANTS .……………………………………………………… 23
2.8.1 VACUOLAR COMPARTMENTALIZATION OF ARSENIC IN PLANTS ………………… 24
2.8.2 ARSENIC CHELATION BY PHYTOCHELATINS ……………………………………….... 25
2.8.3 ARSENIC CHELATION BY GLUTATHIONE .……………………………………………… 26
2.8.4 ANTIOXIDANT ENZYMES AND ARSENIC TOXICITY .………………………………….. 27
2.8.5 SALICYLIC ACID AND ARSENIC TOXICITY .…………………………………………….. 29
2.9 HORMETIC EFFECTS OF ARSENIC TOXICITY .……………………………………………… 29
2.10 CONCLUSIONS AND PERSPECTIVES ………………………………………………………. 30
BIOGEOCHEMICAL BEHAVIOUR OF LEAD IN SOIL-PLANT SYSTEM ……………………... 31
3.1 INTRODUCTION …………………………………………………………………………………… 31
3.2 GLOBAL USES OF LEAD .………………………………………………………………………… 32
3.3 LEVELS AND SOURCES OF LEAD IN SOILS ……………………………………………….... 33
3.4 PHYTOAVAILABILITY OF LEAD IN SOILS …………………………………………………….. 34
3.5 SOIL PROPERTIES AND LEAD PHYTOAVAILABILITY .……………………………………… 35
3.5.1 SOIL PH AND LEAD PHYTOAVAILABILITY …………………………………………….... 35
3.5.2 SOIL ORGANIC MATTER AND LEAD PHYTOAVAILABILITY .………………………… 37
3.5.3 SOIL MICROBIAL ACTIVITY AND LEAD PHYTOAVAILABILITY .……………………... 37
3.6 SOIL-PLANT TRANSFER OF LEAD .…………………………………………………………… 38
3.6.1 MOLECULAR UNDERSTANDING OF LEAD BIO-ABSORPTION …………………….. 38
3.6.2 LEAD SEQUESTRATION IN PLANT ROOTS ……………………………………………. 38
3.6.3 LEAD TRANSLOCATION FROM ROOTS TO SHOOTS .……………………………….. 39
3.7 TOXIC EFFECTS OF LEAD .…………………………………………………………………….. 40
3.7.1 LEAD TOXICITY TO PLANT GROWTH .………………………………………………….. 40
3.7.2 LEAD GENOTOXICITY .…………………………………………………………………….. 41
3.7.3 LEAD-INDUCED OXIDATIVE STRESS ..………………………………………………….. 42
3.8 LEAD DETOXIFICATION ..……………………………………………………………………….. 44
3.8.1 VACUOLAR COMPARTMENTALIZATION OF LEAD ..………………………………….. 44
3.8.2 LEAD CHELATION BY PHYTOCHELATINS ..……………………………………………. 44
3.8.3 LEAD CHELATION BY GLUTATHIONE ..…………………………………………………. 46
3.8.4 ANTIOXIDANT ENZYMES AND LEADTOXICITY ...……………………………………… 46
3.8.5 SALICYLIC ACID AND LEAD TOXICITY ..………………………………………………… 48
3.9 HORMETIC EFFECT OF LEAD TOXICITY ..…………………………………………………… 49
3.10 CONCLUSIONS ..………………………………………………………………………………… 50
BIOGEOCHEMICALBEHAVIOUR OF MERCURY IN SOIL-PLANT SYSTEM ………………… 51
4.1 INTRODUCTION ..…………………………………………………………………………………. 51
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4.2 GLOBAL USES OF MERCURY………………………………………………………………….. 52
4.3 MERCURY LEVELS AND SOURCES IN SOIL………………………………………………… 53
4.4 PHYTOAVAILABILITY OF MERCURY………………………………………………………….. 54
4.5 SOIL PROPERTIES AND HG BIO-AVAILABILITY…………………………………………….. 55
4.5.1 SOIL PH AND MERCURY PHYTOAVAILABILITY ..……………………………………... 55
4.5.2 SOIL ORGANIC MATTER AND MERCURY PHYTOAVAILABILITY …………………… 56
4.5.3 SOIL MICROBIAL ACTIVITY AND MERCURY BEHAVIOUR…………………………… 56
4.6 SOIL-PLANT TRANSFER OF MERCURY……………………………………………………… 57
4.6.1 MOLECULAR UNDERSTANDING OF MERCURY ABSORPTION BY PLANTS ..…… 57
4.6.2 COMPETITION BETWEEN MERCURY AND CATIONS FOR ABSORPTION ..……… 58
4.6.3 MERCURY SEQUESTRATION INTO PLANT ROOTS ………………………………….. 59
4.6.4 MERCURY TRANSLOCATION TO PLANT SHOOTS …………………………………… 59
4.6.5 MERCURY ACCUMULATION IN EDIBLE PLANT PARTS ……………………………… 60
4.7 EFFECTS OF MERCURY ON PLANTS ………………………………………………………… 60
4.7.1 MERCURY TOXICITY TO PLANT GROWTH .…………………………………………… 61
4.7.2 MERCURY GENOTOXICITY ………………………………………………………………. 62
4.7.3 MERCURY-INDUCED OXIDATIVE STRESS …………………………………………….. 63
4.8 MERCURY DETOXIFICATION IN PLANTS …………………………………………………… 65
4.8.1 VACUOLAR COMPARTMENTALIZATION OF MERCURY …………………………….. 65
4.8.2 MERCURY CHELATION BY PHYTOCHELATINS ………………………………………. 65
4.8.3 MERCURY CHELATION BY GLUTATHIONE …………………………………………….. 65
4.8.4 ANTIOXIDANT ENZYMES AND MERCURY TOXICITY ………………………………… 66
4.8.5 SALICYLIC ACID AND MERCURYTOXICITY ……………………………………………. 68
4.9 CONCLUSIONS AND PERSPECTIVES ..……………………………………………………… 68
BIOGEOCHEMICAL BEHAVIOUR OF COBALT IN SOIL-PLANT SYSTEM …………………. 71
5.1 INTRODUCTION ………………………………………………………………………………….. 71
5.2 GLOBAL USES OF COBALT ……………………………………………………………………. 72
5.3 COBALT LEVELS AND SOURCES IN SOILS ………………………………………………… 72
5.4 PHYTOAVAILABILITY OF COBALT IN SOILS ………………………………………………… 73
5.5 SOIL PROPERTIES ADN COBALT PHYTOAVAILABILITY …………………………………. 73
5.6 SOIL - PLANT TRANSFER OF COBALT ………………………………………………………. 74
5.6.1 MOLECULAR UNDERSTANDING OF COBALT ABSORPTION BY PLANTS ……….. 74
5.6.2 COMPETITION BETWEEN COBALT AND CATIONS FOR ABSORPTION ………….. 75
5.6.3 COBALT SEQUESTRATION INTO PLANT ROOTS …………………………………….. 75
5.6.4 COBALT TRANSLOCATION TO SHOOTS ……………………………………………….. 76
5.7 FOLIAR ABSORPTION OF COBALT …………………………………………………………… 76
5.8 TOXIC EFFECTS OF COBALT ON PLANTS ………………………………………………….. 76
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5.8.1 COBALT TOXICITY TO PLANT GROWTH ……………………………………………….. 76
5.8.2 COBALT GENOTOXICITY ………………………………………………………………….. 78
5.8.3 COBALT-INDUCED OXIDATIVE STRESS AND LIPID PEROXIDATION …………….. 79
5.9 COBALT DETOXIFICATION IN PLANTS ……………………………………………………… .80
5.9.1 VACUOLAR COMPARTMENTALIZATION OF COBALT ……………………………….. 80
5.9.2 COBALT CHELATION BY PHYTOCHELATINS ………………………………………….. 80
5.9.3 COBALT CHELATION BY GLUTATHIONE ………………………………………………. 81
5.9.4 ANTIOXIDANT ENZYMES AND COBALT TOXICITY …………………………………… 81
5.10 HORMETIC EFFECT OF COBALT TOXICITY ……………………………………………….. 82
5.11 CONCLUSIONS AND PERSPECTIVES ……………………………………………………… 83
BIOGEOCHEMICAL BEHAVIOUR OF NICKEL IN SOIL-PLANT SYSTEM …………………… 85
6.1 INTRODUCTION ………………………………………………………………………………….. 85
6.2 GLOBAL USES OF NICKEL ……………………………………………………………………… 86
6.3 NICKEL LEVELS AND SOURCES IN SOILS ………………………………………………….. 87
6.4 PHYTOAVAILABILITY OF NICKEL IN SOILS ………………………………………………… 88
6.5 SOIL PROPERTIES AND NI PHYTOAVAILABILITY …………………………………………. 88
6.5.1 SOIL PH AND NI PHYTOAVAILABILITY ………………………………………………….. 88
6.5.2 SOIL ORGANIC MATTER AND NI PHYTOAVAILABILITY ……………………………… 89
6.6 SOIL-PLANT TRANSFER OF NICKEL …………………………………………………………. 89
6.6.1 MOLECULAR UNDERSTANDING OF NI ABSORPTION BY PLANTS ……………….. 89
6.6.2 COMPETITION AMONG NI AND CATIONS FOR ABSORPTION ……………………… 90
6.6.3 NICKEL TRANSLOCATION FROM ROOT TO SHOOTS ……………………………….. 91
6.7 TOXIC EFFECTS OF NI ON PLANTS ………………………………………………………….. 91
6.7.1 NICKEL TOXICITY TO PLANT GROWTH ………………………………………………… 92
6.7.2 NICKEL GENOTOXICITY …………………………………………………………………… 92
6.7.3 NICKEL-INDUCED OXIDATIVE STRESS ………………………………………………… 93
6.8 NICKEL DETOXIFICATION IN PLANTS ……………………………………………………….. 93
6.8.1 VACUOLAR COMPARTMENTALIZATION OF NI ……………………………………….. 93
6.8.2 NICKEL CHELATION BY PHYTOCHELATINS …………………………………………… 95
6.8.3 NICKEL CHELATION BY GLUTATHIONE ………………………………………………… 95
6.8.4 NICKEL TOXICITY AND SALICYLIC ACID ………………………………………………. 95
6.8.5 ANTIOXIDANT ENZYMES AND NI TOXICITY …………………………………………… 96
6.9 HORMETIC EFFECT OF NITOXICITY …………………………………………………………. 97
6.10 CONCLUSIONS AND PERSPECTIVES ………………………………………………………. 98
BIOGEOCHEMICAL BEHAVIOUR OF ZINC IN SOIL-PLANT SYSTEM ………………………. 99
7.1 INTRODUCTION ………………………………………………………………………………….. 99
7.2 GLOBAL USES OF ZINC ……………………………………………………………………….. 100
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7.3 ZINC LEVELS AND SOURCES IN SOILS ……………………………………………………. 101
7.4 SOIL PROPERTIES AND ZINC PHYTOAVAILABILITY …………………………………….. 102
7.4.1 RHIZOSPHERE AND ZINC PHYTOAVAILABILITY ……………………………………. 103
7.4.2 MICROBIAL ACTIVITY AND ZINC BEHAVIOUR IN SOILS …………………………… 104
7.5 SOIL-PLANT TRANSFER OF ZINC …………………………………………………………… 104
7.5.1 MOLECULAR UNDERSTANDING OF ZINC ABSORPTION BY PLANTS ………….. 104
7.5.2 COMPETITION BETWEEN ZINC AND OTHER CATIONS FOR ABSORPTION …… 105
7.5.3 ZINC SEQUESTRATION INTO PLANT ROOTS ……………………………………….. 105
7.5.4 ZINC TRANSLOCATION INTO PLANT SHOOTS ……………………………………… 106
7.5.5 FOLIAR ABSORPTION OF ZINC ………………………………………………………… 106
7.6 ROLES OF ZINC IN PLANTS ………………………………………………………………….. 106
7.6.1 ROLE OF ZINC IN ENZYMES ……………………………………………………………. 107
A. ALCOHOL DEHYDROGENASE ………………………………………………………….. 107
B. CARBONIC ANHYDRASE ………………………………………………………………… 107
C. CUZN-SUPEROXIDE DISMUTASE ……………………………………………………… 108
D. OTHER ZN-CONTAINING ENZYMES …………………………………………………... 108
7.6.2 ZINC AND PROTEIN SYNTHESIS ………………………………………………………. 108
7.6.3 ZINC AND MEMBRANE STABILITY ……………………………………………………... 108
7.7 ZINC DEFICIENCY ……………………………………………………………………………… 109
7.8 ZINC TOXICITY ………………………………………………………………………………….. 110
7.8.1 ZINC TOXICITY TO PLANT GROWTH ………………………………………………….. 110
7.8.2 ZINC-INDUCED OXIDATIVE STRESS …………………………………………………... 110
7.9 ZINC DETOXIFICATION IN PLANTS ………………………………………………………….. 112
7.9.1 VACUOLAR COMPARTMENTALIZATION OF ZINC …………………………………… 112
7.9.2 ZINC CHELATION BY PHYTOCHELATINS …………………………………………….. 112
7.9.3 ZINC CHELATION BY GLUTATHIONE ………………………………………………….. 113
7.9.4 ANTIOXIDANTS ENZYMES AND ZINC TOXICITY …………………………………….. 114
7.10 CONCLUSIONS AND PERSPECTIVES …………………………………………………….. 116
REFERENCES ……………………………………………………………………………………….. 117
SUBJECT INDEX …………………………………………………………………………………….. 189
GLOSSARY ..………………………………………………………………………………………... 194
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LIST OF TABLES
TABLE 1.1: SOME BASIC CHARACTERISTICS OF HEAVY METALS AND METALLOIDS ............................... 6
TABLE 1.2: NATURAL AND ANTHROPOGENIC SOURCES OF SELECTED METALS AND METALLOIDS………7
TABLE 2.1: ARSENIC-INDUCED ENHANCED PRODUCTION OF ROS IN VARIOUS PLANT SPECIES GROWN
IN HYDROPONIC CULTURE ............................................................................................................. 24
TABLE 2.2: ARSENIC-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONIC CULTURE ................................................................................................. 28
TABLE 3.1: LEAD CONCENTRATIONS IN DIFFERENT CONSTITUENTS OF EARTH'S CRUST .................... 34
TABLE 3.2: LEAD-INDUCED ENHANCED PRODUCTION OF ROS IN VARIOUS PLANT SPECIES GROWN IN
HYDROPONIC CONDITIONS ............................................................................................................. 43
TABLE 3.3: LEAD-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONICS……………………………………………………………………..............45
TABLE 4.1: MERCURY CONTENTS IN DIFFERENT SOIL ORDERS AND ROCK TYPES.............................. 54
TABLE 4.2: MERCURY-INDUCED ENHANCED PRODUCTION OF ROS IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONICS .............................................................................................................. 64
TABLE 4.3: MERCURY-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONICS .............................................................................................................. 67
TABLE 5.1: COBALT-INDUCED ENHANCED PRODUCTION OF ROS IN VARIOUS PLANT SPECIES GROWN IN
HYDROPONICS .............................................................................................................................. 79
TABLE 5.2: COBALT-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONIC ................................................................................................................ 82
TABLE 6.1: NICKEL-INDUCED ENHANCED PRODUCTION OF ROS IN VARIOUS PLANT SPECIES GROWN IN
HYDROPONIC ................................................................................................................................ 94
TABLE 6.2: NICKEL-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONIC ................................................................................................................ 97
TABLE 7.1: ZINC-INDUCED ENHANCED PRODUCTION OF ROS IN DIFFERENT PLANT SPECIES GROWN IN
HYDROPONIC .............................................................................................................................. 113
TABLE 7.2: ZINC-INDUCED ACTIVATION OF ANTIOXIDANT ENZYMES IN DIFFERENT PLANT SPECIES
GROWN IN HYDROPONIC .............................................................................................................. 115
ANNEXURE TABLE 1: HARMFUL EFFECTS OF ARSENIC ON PLANT GROWTH .................................... 179
ANNEXURE TABLE 2: HARMFUL EFFECTS OF LEAD ON PLANT GROWTH .......................................... 181
ANNEXURE TABLE 3: HARMFUL EFFECTS OF MERCURY ON PLANT GROWTH ................................... 183
ANNEXURE TABLE 4: HARMFUL EFFECTS OF COBALT ON PLANT GROWTH ...................................... 185
ANNEXURE TABLE 5: HARMFUL EFFECTS OF NICKEL ON PLANT GROWTH ....................................... 186
ANNEXURE TABLE 6: HARMFUL EFFECTS OF ZINC ON PLANT GROWTH........................................... 187
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LIST OF FIGURES
FIGURE 1.1: PERIODIC TABLE HIGHLIGHTING SELECTED HEAVY METAL(LOID)S (SHADED CELLS) ……...3
FIGURE 1.2: BIOGEOCHEMICAL BEHAVIOUR OF HEAVY METALS IN SOIL-PLANT SYSTEM........................9
FIGURE 2.1: ANNUAL WORLD REFINERY PRODUCTION OF ARSENIC (US-GS 2016). ......................... 12
FIGURE 2.2: BIOGEOCHEMICAL BEHAVIOUR OF ARSENIC IN SOIL-PLANT SYSTEM. ............................. 13
FIGURE 2.3: VACUOLAR COMPARTMENTATION OF ARSENIC IN PLANT CELLS. .................................... 23
FIGURE 3.1: ANNUAL WORLD REFINERY PRODUCTION OF LEAD (USGS 2016). ................................ 33
FIGURE 3.2: BIOGEOCHEMICAL BEHAVIOUR OF LEAD IN SOIL-PLANT SYSTEM .................................... 36
FIGURE 3.3: TOXIC EFFECTS OF LEAD ACCUMULATION IN PLANTS. ................................................... 40
FIGURE 3.4: LEAD-INDUCED OXIDATIVE STRESS IN PLANT CELLS ..................................................... 45
FIGURE 3.5: STRAIGHT AND INVERTED U-SHAPED HERMETIC CURVES. ............................................ 49
FIGURE 4.1: ANNUAL WORLD REFINERY PRODUCTION OF MERCURY (USGS 2016). ......................... 52
FIGURE 4.2: MERCURY ABSORPTION AND TRANSFORMATION IN PLANTS. ......................................... 58
FIGURE 5.1: ANNUAL WORLD REFINERY PRODUCTION OF COBALT (USGS 2016). ............................ 72
FIGURE 5.2: BIOGEOCHEMICAL BEHAVIOUR OF COBALT IN SOIL-PLANT SYSTEM. ............................... 77
FIGURE 6.1: ANNUAL WORLD REFINERY PRODUCTION OF NICKEL (US-GS 2016). ............................ 87
FIGURE 6.2: BIOGEOCHEMICAL BEHAVIOUR OF NICKEL IN SOIL-PLANT SYSTEM. ................................ 90
FIGURE 7.1 ANNUAL WORLD REFINERY PRODUCTION OF ZINC (USGS 2016). ................................ 100
FIGURE 7.2 . ENTRY OF ZINC INTO PLANT CELL WALLS AND ITS DETOXIFICATION MECHANISMS. ....... 111
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LIST OF ABBREVIATIONS
½O2 Singlet Oxygen
ABC transporters ATP-Binding Cassette Transporters
APX Ascorbate Peroxidase
As Arsenic
ASC Ascorbate
ATP Adenosine Triphosphate
ATSDR Agency for Toxic Substances and Disease Registry
CAT Catalase
CCA Chromated Copper Arsenate
Cd Cadmium
CDF Cation Diffusion Facilitators
Co Cobalt
DHAR Dehydroascorbate Reductase
DMAA Dimethyl Arsenic Acid
DMT Divalent Metal ion Transporter
DOM Dissolved Organic Matter
EEA European Environmental Agency
Eh Reduction Potential
EPA Environmental Protection Agency
FOREGS Forum of European Geological Surveys
GPX Guaiacol Peroxidise
GR Glutathione Reductase
GSH Glutathione or Reduced Glutathione
GSSG Oxidized Glutathione
H2O2 Hydrogen Peroxide
Hg Mercury
HO• Hydroxyl
IARC International Agency for Research on Cancer
JECFA Joint FAO/WHO Expert Committee Food Additives
LOX Lipoxygenase
MeHg Methylmercury
MDHAR Monodehydroascorbate Reductase
MMAA Mono Methyl Arsenic Acid
Mn Manganese
Mo Molybdenum
NADPH Nicotinamide adenine dinucleotide phosphate
Ni Nickel
NIP Nodulin-26-like Intrinsic Proteins
NRAMP Natural Resistance-Associated MacroPhage
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O2• − Superoxide Anion
OM Organic matter
P Phosphorous
Pb Lead
PC Phytochelatins
PCS Phytochelatin synthase
PGPR Plant Growth-Promoting Bacteria
POD Peroxidase
RAPD Random Amplified Polymorphic
RO• Alkoxyl
RO2• Peroxyl
ROOH Organic Hydroperoxide
ROS Reactive Oxygen Species
SA Salicylic Acid
Sb Antimony
SCE Sister Chromatid Exchanges
Se Selenium
SOD Superoxide Dismutase
SOM Soil Organic Matter
TBARS Thiobarbituric Acid Reactive Substances
TMA Trimethylarsine Oxide
USGS United States Geological Survey
ZIP ZRT, IRT-like Protein
Zn Zinc
Zr Zirconium
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PREFACE
hirteen years of extensive research work on biogeochemical behaviour of heavy metals in soil-plant system, and their potential ecotoxic effect in plants as well as associated human health risks is summarized in this book. This year’s long research journey started with my
work on biogeochemical behaviour of nickel under the visionary supervision of Prof. Dr. Abdul Ghafoor during my M.Sc. (Hons.) in Soil Science at Institute of Soil and Environmental Sciences, University of Agriculture Faisalabad, Pakistan. This journey continued as I did my MPhil (2006-2007), PhD (2007-2010) and Post-PhD (2010-2011) research work from INP-ENSATS, Toulouse-France under world-renowned scientists: Prof. Dr. Camille Dumat and Prof. Dr. Eric Pinelli. My Ph.D. research work was about biogeochemical behaviour of heavy metals in soil-plant system, which was funded by HEC-Pakistan under “Overseas Scholarship Scheme”. I served as a lecturer for three years (2007-2010) at INP-ENSATS, Toulouse, France.
One book, nine book chapters and >52 impact factor articles in scientific literature regarding biogeochemical behaviour of heavy metals in soil-plant system with an accumulative impact factor >125 and citations >1250 precedes this book which I contributed in collaboration with my supervisors and other scientists from different countries. I have supervised (PI & Co-PI) > 25 MPhil students of Environmental Sciences which also broadened my exposure and experience in my research journey. I served as supervisory committee member of three PhDs and one Post-Doc researcher. Based on quality of research work and publications, my doctoral thesis was awarded Leopold Escande Award-2011 by INP-ENSAT France. I also received four “Research Productivity Awards” (2013-14, 2014-15, 2015-16 & 2016-17) by Pakistan Council for Science & Technology (PCST), and five “Research Productivity Awards” (2012, 2013, 2014, 2015 & 2016) by COMSATS Institute of Information Technology based on my research productivity that increased my morale and provided me enough motivation to move further in my extensive research journey.
This book is the collection of scientific data compiled on the basis of my thirteen years of research experience. The valuable information regarding biogeochemical behaviour of nickel, lead, arsenic, cobalt, mercury and zinc in soil-plant system on the same pattern makes this book a worth reading manuscript. Therefore, this book can be a useful window of information for the readers to compare the biogeochemical behaviour of heavy metals in soil-plant system.
Shahid
T
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FOREWORD
Our planet “Earth” is in danger and its morphology is getting worse and worse every day.
Environmental contamination by heavy metals through natural processes and anthropogenic activities is a prevalent and severe dilemma of society, and it is getting significant attention by researchers and regulatory authorities around the globe. Although, the origins of heavy metals contamination date back to ancient times, the dilemma became more profound after the rise of the industrial empires. This industrial revolution owes to the impact of modern technologies and enhanced metal production.
Numerous studies have validated this phenomenon that soil contamination by heavy metals results in the accumulation and uptake of these metals in crops that not only decrease crop growth and productivity but also affect animal and human health. This book is another valuable contribution to the scientific literature as it provides comprehensive knowledge regarding Biogeochemical Behaviour of Heavy Metals in soil-plant system. So, it is no less than a saviour because biogeochemical behaviour (bioavailability in soil, absorption, uptake and accumulation by plants, and toxicity and detoxification in plants) of heavy metals in soil-plant system is considered as a serious environmental concern and this book is providing a holistic knowledge about it.
Seven chapters in this book present the data available regarding biogeochemical behaviour of six heavy metals, i.e., arsenic (As), mercury (Hg), lead (Pb), cobalt (Co), nickel (Ni) and zinc (Zn) in soil-plant system in an easy and understandable manner. Heavy metals behaviour in soil-plant system (mobility, bioavailability, speciation, soil to plant transfer, toxicity and detoxification) differs greatly with respect to plants and metal type, and physico-chemical properties of soils. Six heavy metals selected to be discussed are the most toxic ones classified as human carcinogens according to the Agency for Toxic Substances and Disease Registry [ATSDR 2015a).
In this book, I have summarised data regarding: (i) bioavailability of selected metals in soil, (ii) uptake of metals by plants, (iii) factors affecting metals bioavailability in soil and uptake by plants, (iv) heavy metals compartmentation in different plant organs, (v) toxic effect of metals on plants, and (vi) detoxification mechanisms adopted by plants against the toxic effects induced by these metals.
Shahid
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ACKNOWLEDGEMENT
Without the grace and the power of Allah Almighty, this thirteen years long and arduous
journey would not have been successfully ended into this valuable addition of knowledge into scientific literature. This journey, though unconventional and sometimes disrupted, has always been driven by the light bestowed on our path by the Holy Prophet Muhammad (P.B.U.H.).
I am indebted to Dr. S.M. Junaid Zaidi, Rector CIIT and Prof. Khair-uz-Zaman, Director CIIT-Vehari Campus for providing favourable working environment that we enjoy while working at COMSATS Institute of Information Technology, Vehari Campus. I also thankful to the officials of ORIC-CIIT for their support during the entire period of book writing.
I feel enormously grateful and humbled by all of the amazing people who have guided, inspired and supported me throughout this publication. This book would not have been possible without the major contribution of Sana Khalid, Behzad Murtaza, Hafiz Faiq Siddique Gul Bakhat, Ghulam Abbas, Saliha Shamshad, Marina Rafiq, Sunaina Abbas, Zahida Zia, Faryal Naeem, Nabeel Khan Niazi, Muhammad Imtiaz Rashid, Ghulam Mustafa Shah, Rizwan Ashraf, Hafiz Mohkum Hammad, Muhammad Imran, Irshad Bibi, Muhammad Amjad and Natasha. Finding words comprehensive enough to express my gratitude to all these wonderful people are difficult indeed. I am grateful to all these scientists whose contributions and recommendations made this book a valuable and quality piece of work.
A special thanks goes to my family for their unwavering love, constant support and encouragement to realize my goals. They kept their faith in my abilities to complete this mission. A distinct love to my princes, Asna, who is the most beautiful gift and blessing of God to me in this world.
I would also like to acknowledge the invaluable support of Higher Education Commission (HEC), Pakistan for providing us opportunity to publish this book and financial support. The continuous support, guidance and involvement of HEC officials throughout this process made this project finalized.
Shahid
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CHAPTER-1
POLLUTION OF HEAVY METALS AND
METALLOIDS
Soil contamination with potentially eco-toxic and persistent heavy metals and metalloids are ubiquitous around the world. Concentration of such metals in soils has increased considerably over the past three-four decades, thereby conferring serious risk to the environment and human health (Goix et al. 2015; Sabir et al. 2015; Khalid et al. 2017a). Contrasting to organic contaminants, heavy metals are somewhat unique by the fact that these are highly resistant to either biologically or chemically-induced degradation. Therefore, total heavy metals contents in soils persist for a long duration after being introduced into soils, thereby, present serious environmental concerns.
Contamination of soil by persistent heavy metals results in uptake and accumulation of these metals in crops that not only decrease crop productivity but endure a serious risk to animal and human health (Shahid et al. 2015a; Xiong et al. 2016a). Excessive heavy metals accumulation in plant tissues is capable of provoking a wide range of biochemical, morphological and physiological deleterious effects (Shahid et al. 2015a). Heavy metals mediated increased production of reactive oxygen species (ROS) is reported to be main cause of their toxicity in plants (Flora 2011; Begovic et al. 2016). Heavy metals-induced toxicity to normal biological rocesses/metabolism can be either passive or because of their resemblance with essential ions such as iron (Fe), manganese (Mn), potassium (K) etc. (Pourrut et al. 2011). In order to mitigate toxic effects of heavy metals, plants have developed several defence mechanisms. These defence mechanisms assist plants to maintain crop productivity and sustain the cellular redox state by mitigating the toxicity provoked by metal-induced oxidative stress (Shahid et al. 2014a; López-Moreno et al. 2016).
Studies have shown that impacts of toxic metals on human and environmental health are very bad and many times fatal. For example, when humans are exposed to high levels of toxic metals or metalloids, these affect circulatory functions by disturbing the normal functioning of kidneys, lungs, liver, digestive and nervous system and by affecting blood composition (Järup 2003). Though, the toxicity of metals to living organisms has been known since long, exposure to metals is still a serious environmental and health concern and is even becoming more distinct and severe in some areas of the world. The production and mining of heavy metals and metalloids increased sharply for > 100 years since the middle of the 19thcentury. Therefore, it is dire need of the time to pay more attention on the biogeochemical behaviour (mobility, bioavailability, uptake, compartmentation, toxicity and detoxification of chemical elements and compounds) and fate of heavy metals in the environment. It is necessary to understand links among heavy metal adsorption/ desorption in soils, phytoavailability, and compartmentation inside plants, phytotoxicity and detoxification. Moreover, improved understanding of these mechanisms is necessary in various research areas and remediation techniques such as phytoremediation or choice of traits and plant species in genetic engineering or other plant breeding programs. In this book, I intend to propose relationships that exist between heavy metals and metalloids bioavailability in soils, phyto-uptake, accumulation, translocation, toxicity and detoxification. This book presents the data regarding biogeochemical behaviour of six heavy metals, i.e. arsenic (As), mercury (Hg), lead (Pb), cobalt (Co), nickel (Ni) and zinc (Zn) in soil-plant system (Fig. 1.1).
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1.1 METALS AND TYPES
Metal is a substance which is typically shiny, hard, ductile, malleable and fusible with good electrical and thermal conductivity. There exist around 91 metals out of the 118 identified chemical elements in the Periodic Table. All metals are solid under ordinary conditions except Hg. Researchers have used different terms of metals in the scientific literature such as metalloids, essential metals, light metals, precious metals, heavy metalloids, trace metals, rare earth metals and toxic metals. Several definitions have been proposed and used to describe these groups of metals.
A metalloid is an element that has properties of both the metals and the non-metals such as As. Metalloids generally behave chemically like a non-metal but possess physical properties and appearance of a metal.
Essential metals are named as these play a key role in the metabolic or biochemical process of living organisms and must be present between certain ranges at which they are not toxic to living organisms. These essential metals provide vital co-factors for enzymes and metalloproteins and, thus, their deficiency may seriously and adversely affect metabolic activities. Essential metals include manganese (Mn), zinc (Zn), iron (Fe), copper (Cu), magnesium (Mg), molybdenum (Mo), chromium (Cr) and selenium (Se) (Maret et al. 2016).
Trace metals generally refers to those metals which naturally occur at very low concentration (< 1 mol m-3) in soils. Trace metals include Mg, Fe, Zn, Li, Cr, Cu, Co, Ni, Se, Mn and others.
Toxic metals are those which induce and provoke toxicity to living organisms even at very low concentrations. These metals negatively affect plant metabolism by altering the dynamic conformation of biological molecules, replacing other metal ions and blocking the vital functional groups. Example includes Pb, As, Cd, Hg etc.
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Figure 1.1: Periodic table highlighting selected heavy metals (shaded cells) and metalloids
25
Precious metals are generally rare in the Earth crust with high economic value. These metals are generally less reactive, ductile and have a high lustre than most of the elements. Platinum (Pt), gold (Au), silver (Ag) and palladium (Pd) are the best known precious metals.
Light metals have an elemental density 5 g cm-3. Aluminium, Mg and Ti are the light metals having sufficient commercial significance.
Rare earth metals are a group of chemically similar metallic elements comprising the lanthanide series. These metals generally occur together in nature and are difficult to distinguish from one another.
Heavy metals are naturally occurring elements that have a density > 5 times higher than water.
1.2 HEAVY METALS AND METALLOIDS
The term “heavy metal” has been cited increasingly in publications and in legislation related to risk assessment and remediation studies. This term is often used as a group name for metals and metalloids generally linked with toxicity and risk assessment, environmental contamination and remediation. Therefore, the lists of heavy metals specified by various legal regulations/organisations may vary from one another, or the terms may be used without describing which metalloids are covered. In other words, the term “heavy metal” has been used contradictorily.
According to Garbisu and Alkorta (2003), this term is vague and arbitrary. Several definitions of heavy metals in the past have been proposed or published over time. Some authors described heavy metal as any element having an atomic number greater than 20 and possessing the properties of a metal, like density, conductivity, ductility, ligand specificity, stability as cations, etc. (Raskin et al. 1994).
The most widely accepted definition of heavy metals is based upon the relative high density of a metal. However, the classification of heavy metals based on their elemental densities or specific gravity varied over years. Different authors during different periods defined heavy metals using different values of specific gravity: > 7 (Bjerrum 1936), > 6 (Thornton 1995), > 5 (Parker 1989; Lozet and Mathieu 1991; Morris 1992), > 4.5 (Streit 1994) and >3.5 (Falbe and Regitz 1996). Recently, Park et al. (2011) defined heavy metals as the elements with an atomic density > 6 g cm−3 [excluding As, B and Se]. Some authors suggested that any toxic metal may be considered as heavy metal, regardless of its density and atomic mass. Any metal may be regarded a “contaminant/pollutant” if it exists where it is not wanted, or in a concentration or chemical form that is hazardous to environment or human beings. Heavy metals generally include some metalloids, transition metals, actinides and lanthanides that can induce toxicity to living organisms even at lower levels of exposure. In this book, the term “heavy metals” will be used for potentially toxic elements to plants at exorbitant concentrations.
Heavy metals are natural constituents of the Earth crust. Heavy metals exist at various concentrations in different soil profiles (Kabata-Pendias 2011). It has been reported that out of 91 metals, 53 are regarded as heavy metals and most of these do not play any significant biological function in the metabolic or physiological mechanism of living organisms (Hayat et al. 2010), especially plants. However, Zn, Cu, Se, Co, Ni, Mn, and Mo play key biological roles, and have positive biochemical effects on plant growth and development (Kabata-Pendias 2011; Shahid et al. 2015a). While, a few of these metals do not have any metabolic or biochemical function such as As, Hg, Pb, Cd, Sb and Zr rather may decrease crop quality and productivity when their level increases up to a certain value (Kabata-Pendias 2011; Shahid et al. 2013a). Member metals of both classes cause toxicity even at relatively low amounts, while the former play an important role in regulating biological processes such as catalytic co-factors in various enzymes and electron-transferring proteins. Some heavy metalloids such as As, Pb, Hg, Cd, Ni and Co are included in the priority list of contaminants by Environmental Protection Agency (EPA 2014), owing to their intense potential and wide spread existence at contaminated sites and their well-known harmful
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effects on human and ecological receptors. Similarly, As, Hg, Cd, Pb, Ni, Co and Zn are considered systemic toxicants by the ATSDR (2015a) based on their wide spread occurrence, potential toxicity and the possibility for human exposure via the food chain (Table 1.1).
1.3 HEAVY METAL SOIL CONTAMINATION
Heavy metals occur naturally in the Earth's crust, but indiscriminate human activities have considerably changed their biochemical balance and geochemical cycles. Heavy metalloids accumulation in soils is mainly originated from various natural and anthropogenic sources (Table 1. 2). Natural sources of heavy metalloids soil contamination include weathering, lithogenesis, erosion, desertification and other geological processes. While, the anthropogenic sources are primarily pertains to mining, smelting, ore processing, atmospheric deposition, vehicle exhaust, land application of industrial discharges, wastes, fertilizers, pesticides and other human activities (Kabata-Pendias and Pendias 2011; Wuana and Okieimen 2011; Schreck et al. 2014). Moreover, accumulation of heavy metals in soils through agricultural practices has also been augmented (Shahid et al. 2014g, 2017a).
Soils are considered the main sink for heavy metals released into the environment by human activities or geochemical transformations. It has been estimated that > 10 million sites are contaminated around the globe, while >50% of these sites are contaminated with heavy metals (He et al. 2015). Compared to developing countries, majority of these heavy metal contaminated sites are reported in developed countries such as the USA, Germany, France, Australia and China because of their excess use in industrial applications (ATSDR 2015a; Goix et al. 2015).
High levels of heavy metal contaminated agricultural sites have been documented near mining sites or natural mineral deposits in different countries of Europe (Shahid et al. 2013b; Goix et al. 2015; Mombo et al. 2016; Khalid et al. 2016). European Environmental Agency (EEA 2007) reported that there are > 2,965,000 potentially polluting activity sites in Europe and about 250,000 heavy metals polluted sites in the EEA member countries. More than 80,000 heavy metals polluted sites have been remediated in Europe during the last 30 years (EEA 2007). Approximately 0.5 million sites in Europe are highly contaminated and need remediation on priority. It is estimated that the number of heavy metals polluted sites demanding clean-up may increase by > 50 % by 2025 (EEA 2007). Large areas have been contaminated with heavy metals in USA alone (USGS 2016). The US EPA has reported > 50,000 priority heavy metal polluted sites in the USA requiring urgent remediation (Ensley 2000). Gorospe (2012) collected soil samples from 91 vegetable gardens in San-Francisco, USA. It was reported that >75% of the gardens showed heavy metal levels above the recommended value of California Human Health Screening Level.
High level of heavy metals and metalloids in agricultural soils is also reported in south Asia such as India, Pakistan and Bangladesh (Hussain et al. 2010; Farid et al. 2015; Rehman et al. 2016; Bhatti et al. 2016). In these countries, the major source of heavy metals in agricultural soil is the use of untreated wastewater (industrial and city wastewater) for crop irrigation (Shahid et al. 2014g; Farid et al. 2015). Numerous studies in India and Pakistan highlighted heavy metal accumulation in soils and crops due to crop irrigation by untreated wastewater (Murtaza et al. 2012; Shahid et al. 2014g). In Pakistan, about 64% of wastewater is directly released into water bodies (rivers and canals) without any primary treatment (Ensink et al. 2004). Similarly, about 30% of wastewater is directly used for crop irrigation of 32,500 ha in Pakistan (Ensink et al. 2004). It is reported that untreated wastewater is used to irrigate approximately 26% of vegetables cultivated in Pakistan (Ensink et al. 2004).
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Table 1.1: Some basic characteristics of heavy metals and metalloids
Name Symbol ATSDR1
ranking* Density2
Year of dis-
covery
Inventor name
Mine production
3 Mineral Occurrence
Arsenic As 1 5.7 1250 Albertus Magnus
46,000 Arsenopyrite
China, Japan, Chile, Mexico, Belgium, Namibia, Philippines
Lead Pb 2 11.34 1817 Friedrich Stromey
5,460 Galena, Anglsite, Cerussite
USA, Canada, Mexico, Australia, Peru, Argentina, Bolivia, South Africa,
Mercury
Hg 3 13.6 3,000 BC
NK4 1,870
Cinnabar, Corderoite, Livingstonite
Chile, Peru, Argentina, Canada, Italy, USA, Mexico, China, Japan
Cobalt Co 51 8.9 1737 Geroge Brandt
112,000
Safflorite, Glaucodot, Skutterudite
China, Norway, Russia, Zimbabwe, Germany, Canada, Morocco, Australia
Nickel Ni 57 8.9 1751 Alex Constedt
2,400,000 Garnierite, Pentlandite
Canada, Russia, Australia, Norway, South Africa, Finland, USA
Zinc Zn 75 7.11 1746 Andreas Marggraf
13,300
Zinc Sulfide, Smithsonite, Hemimorphit,
Peru, Canada, Mexico, Turkey, USA, Australia
1 = Agency for Toxic Substances and Disease Registry, 2= g.cm-3 at 20°C, 3= metric tons (ATSDR 2015a), * Based on based on the toxicity, frequency of occurrence, and the potential for human exposure
1.4 HEAVY METALS OCCURENCE IN SOILS
Heavy metals exist in soils in a number of forms those differ considerably regarding their solubility, mobility, bioavailability and biotoxicity. Heavy metals may occur in soils in different forms, like (i) dissolved, (ii) exchangeable, (iii) structural constituent of minerals and (iv) precipitated as secondary compounds. The portion of heavy metals that exists in dissolved form is attributed as the most important metal fraction because of its high availability to plants (Pourrut et al. 2011). The exchangeable fraction of metals is bound with soil solid particles and is the second most readily phytoavailable fraction of metals (Saifullah et al. 2009). On the other hand, fraction of metals which is an intrinsic constituent of minerals is not phytoavailable (Shahid et al. 2012a). The occurrence of metals/nutrients in soils in different chemical/available forms is usually controlled by adsorption-desorption redox processes and reactions of soils (Hussain et al. 2012; Saifullah et al. 2015). These processes are governed by various soil characteristics such as soil mineralogy, pH, organic matter content, amount and type of other metals, cation-exchange capacity as well as the biological and microbial conditions.
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Majority of the heavy metals are sparingly soluble and mobile in soils with very low phytoavailability. In fact, metals mostly occur in soils in residual forms due to strong binding with biotic/abiotic and organic/inorganic ligands (Uzu et al. 2009; Saifullah et al. 2015). For example, Pb and Cu exist in soil mainly in complexed forms. Therefore, these metals have low mobility and phytoavailability in soils. Soil solution concentration of Pb is generally < 1% of its total contents in soil (Pourrut et al. 2011). On the other hand, some metals such as Zn and Cd mainly occur in soil solution as free ionic form, thereby these metals have high soil-plant transfer index (Shahid et al. 2016). Adsorption of Cu and Pb by soil particles is higher than Ni, Zn and Cd (Shahid et al. 2015a).
After absorption, major portion of metals is stored in plant roots (> 90%). Enhanced accumulation of heavy metals in root cells/tissues seems due to their immobilization/binding by negatively charged cell walls or blockage by endodermis (Pourrut et al. 2011). However, plants differ greatly with respect to their ability to absorb and transfer metals to shoots, which determines their utilization in heavy metals contaminated soils. Generally, hyper-accumulator plant species have high soil-root and root-shoot transfer index (Shahid et al. 2012b).
Table 1.2: Natural and anthropogenic sources of selected metals and metalloids
Name Sources References
Arsenic
volcanic rocks, hydrothermal deposits, marine sedimentary rocks, and fossil fuels, pesticides, insecticides, phosphate containing fertilizers, mining and smelting, industrial processes
Garelick et al. 2008; Wuana and Okieimen 2011; Kabata-Pendias and Pendias 2011
Lead batteries, livestock manures, sewage sludge, inorganic fertilizers, industrial wastes, natural geogenic anomalies, and mining activity
Pourrut et al. 2011; Kabata-Pendias and Pendias 2011
Nickel
waste incineration, fossil fuel burning, batteries, volcanic eruptions, windblown dust, mining, smelting, manufacturing of house hold goods such as stainless steel
Cempel and Nikel 2006;
Wuana and Okieimen 2011; Kabata-Pendias and Pendias 2011
Mercury fossil fuel burning, municipal solid waste incineration and hospital incineration
Issaro et al. 2009; Wuana and Okieimen 2011
Cobalt mining, smelting, batteries, tanneries, electronics, petrochemicals, electroplating and use in textile and pesticide mills
Wuana and Okieimen 2011
Zinc metal production, fossil fuel combustion and waste disposal
Kabata-Pendias and Pendias 2011; Wuana and Okieimen 2011
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1.5 POTENTIAL HAZARDS OF HEAVY METAL CONTAMINATION
1.5.1 Environmental Persistence
In contrast to organic contaminants, heavy metals are highly persistent in the environment. Generally, heavy metals are little degraded biologically or chemically. Therefore, metal contents in soils persist for a long time after being introduced into soils. For example, Pb has a soil persistence period of 150–5,000 years, and has been reported as persistent contaminant in soils for more than 150 years after sewage sludge application (Nandakumar et al. 1995). Similarly, average biological half-life of Cd has been evaluated to about 18 years (Forstner 1995). However, heavy metals are not destroyed/degraded by microbial activity, i.e. no change in the nuclear structure of metals (Nies 1999). Nevertheless, soil microorganisms can transform heavy metals and metalloids from one form to another form by oxidation or reduction processes. These redox transformations of metals may cause an increase or decrease in their mobility, solubility, bioavailability and toxicity.
1.5.2 Potential Phyto-Toxicity
Cultivation of plants on soils contaminated with heavy metals could cause their accumulation or build-up in different plant parts as shown in Fig. 1.2 (Mahdavian et al. 2016). Excessive heavy metals concentration in plant tissues is capable of inducing various physiological, morphological and biochemical toxic effects (Shahid et al. 2015a). Heavy metals induce plant toxicity by disrupting nutrient and water absorption and transport, altering nitrogen metabolism, disrupting the activity of Adenosine Triphosphatase (ATPase), decreasing photosynthesis, interfering with plant growth, dysfunctioning plant photosynthetic apparatus in chloroplasts and causing stomatal closure (Shahid et al. 2015a; Zafari et al. 2016). Heavy metals may also cause visible symptoms of plant injury such as browning of roots, necrosis, chlorosis and leaf rolling (Pourrut et al. 2013) although symptoms may vary for different plant types, age, metal contents and environmental conditions (temperature, day light, humidity, wind velocity etc.).
At the cellular level, excessive heavy metals exposure could cause enhanced production of ROS and chromosomal aberrations (Flora 2011; He et al. 2016). Heavy metals-induced toxicity to normal biochemical and physiological processes could be either passive or because of their resemblance with essential ions such as Ca, Mg, K etc. (Krämer et al. 2007). To survive under heavy metal stress conditions, plants have evolved various natural defence systems. These defensive mechanisms help plants to avoid heavy metal toxicity by scavenging heavy metal induced enhanced production of ROS (Hattab et al. 2016).
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Figure 1.1: Biogeochemical behaviour of heavy etals in soil-plant system
1.5.3 Potential Risk of Metals to Human Health
Food safety is the most important public concern. People are becoming more and more concerned regarding the deleterious effects of heavy metals on human and environmental health. Consumption of food contaminated with heavy metals is considered to be the major pathway (> 90%) of human exposure to metals compared to dermal contact or inhalation (Mombo et al. 2016; Xiong et al. 2016b). Soil is the direct pathway for the accumulation of heavy metals in vegetables and crops via root absorption (Pierart et al. 2015). Excessive build-up of heavy metals in agricultural soils results in increased metal absorption by food crops and vegetables, which in turn may induce serious health risks to human beings (Xiong et al. 2016a).
Heavy metals are reported to cause several disorders in humans including cardiovascular diseases, cancer, cognitive impairment, chronic anaemia, and dysfunctioning and ultimately damage of kidneys, nervous system, brain, skin, and/or bones (Jarup 2003; ATSDR 2015a). Lead and arsenic are capable to induce carcinogenesis, teratogenesis and mutagenesis. High Pb and As concentrations in edible plant parts have been attributed to the occurrence of upper gastrointestinal cancer (Jarup 2003; ATSDR 2015a). Moreover, Pb could also cause improper haemoglobin synthesis, renal and tumor infection, elevated blood pressure and dysfunctioning of reproductive system. Owing to potential toxic effects associated with heavy metal exposure, there is a global concern to comply that the heavy metal contents of agricultural soils and crops cultivated there do not exceed the allowable regulatory limits.
pH, OM, EC, CEC, cationsComplexed or adsorbed
heavy metal(loid)s in soil
(adsorption, precipitation,
complexation & redox
reactions)Metal-oxides and clay particles
Microbial activities
Labile heavy metal(loid)s
in soil solution
Leaching to groundwater
(very little)
tra
nlo
ca
tio
nto
sh
oo
ts
Threat of food chain
contamination
Heavy metal(loid)s sources
(natural & anthropogenic)
Mining, smelting, coal burning,
combustion activities and
volcanic eruption, fertilizers,
pesticides, wastewater, biosolids
and manures, industrial waste
Accumulation
in roots
Affect plant growth &
development
Induce ROS
production and
activation of defense
mechanisms
Heavy metal(loid)s may persist
in soil for several years
Vo
latiliz
atio
n
31
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CHAPTER-2
BIOGEOCHEMICAL BEHAVIOUR OF ARSENIC IN SOIL-PLANT SYSTEM
2.1 INTRODUCTION
Arsenic belongs to transitional reactive metalloids and is ubiquitous in the environment. Among the 98 naturally occurring elements, As contents rank 20th in the Earth's crust and 12th in the human body (ATSDR 2015a). Its abundant occurrence in environmental compartments (water, air, biota, land and aquatic sediments) is a dilemma of the wide spread concern. The European Union has recognized the compounds of Arsenic as "toxic" and "dangerous for the environment" under directive 67/548/EEC. Arsenic has been classified as group 1 carcinogen by the US-EPA and International Agency for Research on Cancer (Rosas-Castor et al. 2014). The Agency for Toxic Substances and Disease Registry (ATSDR 2015a) has ranked this pollutant as No. 1 among the top twenty toxic substances. Acute and chronic exposures to As may cause arsenicosis, which is a common term used for As related health ailments including respiratory diseases, non-pitting edema, diabetes, reproductive disorders, liver and cardiovascular problems, gastro-intestinal, high blood pressure and cancer of skin, liver, bladder, kidney and lungs (ATSDR 2007a; Abdul et al. 2015). Consumption of drinking water and As contaminated food are the major public and environmental health concerns around the globe (Shakoor et al. 2015).
Arsenic is released into the environments both by natural and anthropogenic activities. The most widespread sources of As in soils and waters are natural sources, like volcanic activities, erosion and weathering of rocks and minerals (Chung et al. 2014). Thus, a substantial amount of As could thus be transferred from the contaminated aquifer to different environmental compartments, such as soils and waters (Ali et al. 2003). It has been estimated that natural sources release about 12,000 tons of As every year into the environment, which is almost double compared to anthropogenic releases (Pacyna and Pacyna 2001). In addition to natural sources, use of pesticides, different fertilizers, industrial waste, agricultural chemicals and CCA (copper, chromate and arsenate) wood preservative in different agricultural and industrial activities includes the main anthropogenic sources of As contamination in soil, water and air (Chung et al. 2014). Human-induced release of As into the environment also occurs from heating of Au, Cu, Pb and Co ores from smelters (ATSDR 2015a). It is estimated that human activities release around 50,000 tons of As every year into the environment (Bolan et al. 2014). Arsenic production has decreased by 66% from 59,800 to 3600 thousand metric tons from 2006 to 2015 (Fig. 2.1).
Soil to plant transfer and consumption of As contaminated food are considered to be major pathway (> 90%) of human exposure to the toxic heavy metals (Foucault et al. 2013; Rafiq et al. 2017a,b; Mombo et al. 2017). Environmental contamination with As in soil-plant system poses a severe risk for environmental and human health, thereby affecting thousands of people with As toxicosis globally (Fig. 2. 2) (Saha et al. 1999; Abdul et al. 2015). In general, plant availability and mobility of As in soil depend upon the diverse interplay of mutually dependent biotic and abiotic factors such as pHs, desorption - sorption processes, microbially mediated redox processes and the presence of As adsorbing mineral phases (Fedotov et al. 2012).
Bioaccumulation of As in vegetables and crops cultivated on As polluted soils or irrigated with As contaminated waters has become a major emerging issue (Neidhardt et al. 2012). Arsenic absorption by plants is mainly through roots (Neidhardt et al. 2015), which can be via active pathway (requires energy) or passive pathway (does not require energy). Till date, no As-specific transporter for absorption by plants has been reported (Ghosh et al. 2015). It is reported
33
that As[V] enters the plants through phosphorous (P) cell channels (Lou et al. 2015; Niazi et al. 2017). Excessive level of As in soils is alarming for plants due to numerous physiological, morphological and biochemical disorders. It is shown that As is non-essential for plants and cause toxic effects even at low concentrations (Flora 2011). Arsenic interferes with numerous plant metabolic processes because it is assumed to cause direct effect on stomatal apparatus (Khalid et al. 2016) and can adversely affect plant transpiration rates. Arsenic accumulation inhibits the development of root growth which is highly susceptible to As toxicity. Arsenic stress could cause growth suppression and eventually plant death (Khalid et al. 2016). Arsenic causes oxidative stress and lipid peroxidation due to over-production of reactive oxygen species (ROS) such as H2O2 (Flora 2011).
Figure 2.1: Annual world refinery production of arsenic (US-GS 2016).
Now-a-days, it is known that biogeochemical behaviour of As is governed by its chemical form (Smith et al. 2009; Castillo-Michel et al. 2011). Chemical speciation refers to the existence of a metalloid in different chemical forms. Arsenic speciation is greatly influenced by the ambient conditions and it may occur in various oxidation states: −III, 0, +III, and +V (Castillo-Michel et al. 2011). There are different As forms that exist in the environment, e.g. inorganic [arsenious acid, arsenite, arsine, arsenate or arsenic acids], organic [dimethyl arsenic acid (DMAA) and mono methyl arsenic acid (MMAA)], biological and other forms (Meharg and Hartley-Whitaker 2002). Toxicity of As depends on its different forms and generally inorganic As forms [As(V) and As(III)] are more toxic as compared with organic forms (Khalid et al. 2016). The biological availability, toxicity and transport mechanisms of As depend on its chemical forms (Ruiz-Chancho et al. 2007).
30
35
40
45
50
55
60
65
10
3T
ho
us
an
ds
me
tric
to
ns
Arsenic
34
Figure 2.2: Biogeochemical behaviour of arsenic in soil-plant system.
2.2 GLOBAL USES OF ARSENIC
Owing to the useful properties, As and its compounds are extensively used in various industrial processes. The reported world production of As was 46,000 tons in 2014 (USGS 2016). Elemental As is widely used in manufacturing of alloys mainly with Pb and Cu. Arsine and gallium arsenide are commonly used in electronic industries as well as in manufacturing of semi-conductors (ATSDR 2007a). Compounds of As are used for manufacturing pigments, sheep-dips, poisonous baits and leather preservatives. These are used in pyrotechnics, pharmaceutical substances, antifouling agents in paints, catalysts, dyes, soaps, ceramics and electro-photography. The USA is historically the largest consumer of As worldwide. It has been reported that about 90% of As produced in USA was consumed in manufacturing wood preservatives before 2004. The USA and Canada entered into voluntary agreement in 2003 to ban the use of CCA in residential applications.
Arsenic acid
[HAsO4]2−, [H2AsO4]
−
arsenous acid
Arsinites, methyl arsenic
acid, dimethyl arsenic
acid and arsine
Soil
Toxicity
Decreased
biomass, growth and
chlorophyll synthesis,
DNA breakage,
membrane leakage,
enzyme inactivation,
affect on RUBISCO,
ROS generation
Detoxification
Phytochelatins,
low molecular weight
thiols glutathiones,
vacuolar sequestration,
antioxidant enzymes
(CAT, POX, LOX, SOD,
GPX, APX)
Water
Phosphate
transporters
Tra
nslo
ca
tion
As in roots
As in
shoots
Arsenic sources (natural & anthropogenic)
Mining, copper smelting, coal burning, combustion
activities and volcanic eruption, herbicides, insecticides,
rodenticides, wood preservatives, paint, dyes and
semiconductors
35
In agricultural, As has been used widely in various inputs such as defoliants, insecticides, herbicides, cotton desiccants and soil sterilants. Up to 1970s, approximately 80% of the total mined As was used in pesticide industry (Kabata-Pendias 2011). Demand for As has been declining since 1970s but inorganic salts of As are being used extensively as pesticides (Matschullat 2000). Arsenic had been used to manufacture pesticides such as calcium arsenate (CaAsO4), magnesium arsenate (MgAsO4), Paris green [Cu(CH3COO)2*3Cu(AsO2)2], lead arsenate (PbAsO4) and zinc arsenate (ZnAsO4) during 1880 to 1950 (Merry et al.1983). Owing to potential toxicity of As, its use in pesticides has decreased to around 50%, but organic compounds of As are still used for pesticides manufacturing (Kabata-Pendias 2011). In USA, inorganic and organic pesticides containing As have been used before 1993, but now have been forbidden to decrease As contamination (US-EPA 2009).
2.3 ARSENIC LEVELS AND SOURCES IN SOILS
This metalloid was first reported during the year 1250 by Albertus Magnus (Emsley 2001). Arsenic is the 12th most abundant element in the human body, 14th in sea water and 20th in the Earth crust (0.0001%), (Kalia and Joshi 2009). Arsenic is the natural constituent of numerous minerals (Shakoor et al. 2015). Arsenic in uncontaminated sea water ranges from 1 to 2 μgL-1, while from 5 to15 μgg-1 in marine sediments (Kalia and Khambholja 2015). Mean lithosphere level of As is ≈ 5 mg kg-1 (Mehmood et al. 2009). Natural As levels in uncontaminated soils vary from1 to12 mg kg-1 and depend on the conformation of parent materials from which soils have developed. Different authors reported different natural As levels in uncontaminated soils: 3.8 mg kg-1 (Eriksson 2001), 4.7 mg kg-1 (Kabata-Pendias 2011), 7.2 mg kg-1 (Burt et al. 2003) and 11.6 mg kg-1 (FOREGS 2005). Mehmood et al. (2009) reported that soils As concentration is generally in the range of 5-10 mg kg-1 and As concentration in soil above 20 mg kg-1 is deliberated as high. Much higher concentration of As has been demonstrated in the acid sulphate soils developed from pyritic parent materials. Dudas (1987) attributed high levels of As in the range of 8 to 40 mg As kg-1 in the Canadian acidic sulphate soil to weathering of pyrites in parent material. Based on the nature and intensity of natural and anthropogenic sources, As concentration in soils can vary between <1 to 250,000 mg k-1 (Mahimairaja et al. 2005). Studies have reported high variation in As level among soils and countries owing to variation in anthropogenic activities and soil parent materials. It is reported that the total As contents in uncontaminated natural soils is generally < 10 mg kg-1 (Adriano 2001).
Arsenic is released from the Earth's crust to the environment (water, atmosphere and soil) by various natural and anthropogenic sources. The widest spread sources of As in water and soil are natural ones like volcanic activities, weathering and erosion of rocks and minerals geothermal waters (ATSDR 2007a). The dissolution and desorption of naturally occurring As-rocks and minerals are among the main sources of contaminating groundwater and soil (US-GS 1999). It is reported that As is released from minerals and rocks by numerous processes, but a variation from aerobic to anaerobic conditions causes reduction of As resulting in the generation of high concentrations of dissolved As (USGS 1999). In fact, As rarely occurs as a pure metalloid, but as a constituent of sulfur-bearing minerals. Naturally, As is a major constituent in >200 minerals including elemental As, arsenates, arsenite, arsenides, sulfides and oxides. Arsenic contents of igneous rocks differ significantly up to 100 with an average of 2–3 mg kg−1. Sedimentary rocks differ widely in their As contents; Mn ores contain up to 15,000 mg kg−1of As whereas limestone and sandstone contain very small amount of As (usually < 20 mg kg−1) (Bolan et al. 2014). Similarly, sulfide ores contain As up to 8,000 mg kg−1 (Chilvers and Peterson 1987). Therefore, As contents differ greatly in soils and mostly depend on As contents of parent material (ores and rocks) from which soils developed. However, these minerals are relatively very rare in natural environment. Arsenopyrite (FeAsS) is the most abundantly present ore mineral of As.
Atmospheric deposition plays an important role in geochemical cycle of As (O'Neill 1990). About 60% of atmospheric As flux reported due to low temperature volatilization with the volcanic activity and is regarded as the 2nd most important natural source of As in the environment (Chilvers and Peterson 1987). Arsenic concentration of 27mg g-1 was found in soils contaminated
36
with smelter or mine wastes (EPA 1982). Soil samples taken from mining site in Tamar Valley in the South-west England contain As > 50mg g-1 (Erry et al. 1999). In the vicinity of copper smelters in Montana and Anaconda, As levels ranged from 121 to 236 μg g-1 in the top 2 cm of many soils (Hwang et al. 1997).
2.4 PHYTOAVAILABILITY OF ARSENIC IN SOILS
Phytoavailability of an element in soils refers to the degree to which it can be absorbed by plants. Phytoavailability of an element depends on its binding with soil constituents. Generally exchangeable fraction of any metal presents a better indication of its mobility and phytoavailability in soils. Normally, the portion of phytoavailable As to total As concentration in soil is very low and further soil-plant transfer index even for hyperaccumulators is limited (Niazi et al. 2016). Mostly soil-plant transfer of As depends on its chemical forms. Arsenic can form various organic and inorganic compounds in soils. Arsenic occurs in soils mainly in inorganic form either as arsenate [As(V)] or arsenite [As(III)]. Arsenic predominates commonly in +5 oxidation state under the oxic soil conditions (pH = 5-8; Eh > 200 mV), whereas As[III] is abundant under reduced conditions. Both the forms of inorganic As can be inter-converted by chemical and/or microbial oxidation or reduction as well as the methylation reactions in soils (Khalid et al. 2016). Organic form of As (arseno-betaine) in soil is rapidly demethylated to the dimethyl arsinic acid, which induces the formation of As[V] (Huang et al. 2007). Various studies suggested the existence of methylated As forms such as mono-methyl arsonic acid, dimethyl arsenate, and trimethyl arsine oxide in soils (Tripathi et al. 2012). It is reported that As[V] is less mobile, soluble and bioavailable than As[III], due to its strong binding with soil constituents. Arsenate is more stable than As[III] and adheres extensively and strongly to soil particles such as metal oxides, hydroxides, clays and organic matter. Adsorption of As[V] onto soils is generally enhanced when the net charge is positive.
2.5 EFFECT OF SOIL CHEMICAL PROPERTIES ON ARSENIC PHYTOAVAILABILITY
2.5.1 Soil pH and Arsenic Phytoavailability
Soil pH, extent of the alkalinity or acidity of soil, is one of the parameters that mainly govern As phytoavailability, mobility and partitioning between solid phase, soil solution phase and chemical speciation in soil-plant system (Signes-Pastor et al. 2007). At pH < 5 (Carbonell-Barrachina et al. 1999) or 5.5 (Signes-Pastor et al. 2007), mobility and phytoavailability of As in soil increased as soils became more acidic. Soluble As fraction increases generally at lower pH which increases its phytoavailability in soils. Variation in pH of soils also affects As speciation and consequently its behaviour in soil-plant system. Relative distribution of As[III] and As[V] also varies in soil environment with change in soil pH (Adra et al. 2016).
Increasing soil pH induces As precipitation and adsorption to different metal binding sites of soils constituents (Sauve et al. 2000). In solution, As[V] has pKa’s (acid dissociation constant) of 2, 6.9 and 11.5, whereas, As[III] has pKa of 9.2, 12.1 and 13.4 (Goldberg and Johnston 2001). Around circumneutral pH, HAsO4
2-, H2AsO4- and H3AsO3º species predominate. Under oxidising
conditions, H2AsO4 is the dominant form at lower pH (<6.9) whereas, HAsO42– predominates at
higher pH. Under reduced soil conditions and at pH >9.2, H3AsO30 becomes dominant (Yan et al.
2000). It is reported that at higher soil pH, As co-precipitates with Ca or SO4 (Moreno-Jiménez et al. 2012). Arsenic is mainly sorbed onto Al and Fe oxides under the acidic conditions whereas this form is sorbed onto Ca under alkaline soil conditions (Pongratz 1998). Increased sorption of As[III]) by montmorillonite and kaolinite is also reported over a pH range of 3-9 (Griffin et al. 1978). The H2AsO4
- and HAsO42 are the predominant As forms in soil at pH range of 3-11
(Smedley and Kinniburgh 2002).
Generally, there exists a negative correlation between soil pH and phytoavailability of metals. An increase in soil pH decreased metal absorption and adsorption as well as uptake
37
(Arco-Lazaro et al. 2016). A negative correlation between soil pH and As uptake in Oryza sativa was reported by Bhattacharya et al. (2010). However, Ahmed et al. (2011) reported a positive correlation between soil pH and As phytovailability in Oryza sativa. Arsenic, both As[V] and As[III], has strong affinity for inorganic components of soils and therefore, forms strong bonds with clay minerals and oxides and hydroxides of Al and Fe (Giles et al. 2011). This adsorption of As onto soil constituents is greatly affected by soil pH in addition to other physicochemical properties of soils. For example, Burns et al. (2006) evaluated the adsorption of As on 18 soils with varying soil properties, and described a strong negative correlation between As adsorption and soil pH. An increase in the adsorption of As[V] by goethite with decrease in pH was reported with maximum adsorption at pH 5 (Dimirkou et al. 2009). A sequential fractionation procedure was used to evaluate the effect of change in soil pH (from 4 to 8) on As adsorption - desorption by soil constituents. It was observed that at lower pH (4), Fe-As was the most dominant form followed by Al-As, whereas at higher pH (6-8), Ca-As was predominant form (Akins and Lewis 1976). Similarly, Adriano (2001) revealed that Fe-P and Al-P predominated under acidic soils while Ca-P predominated under calcareous and alkaline soils. In fact, when soil pH is greater than 8, the net negative surface charge of oxides and hydroxides of Al and Fe can deter negatively charged ions of As such as As[V]. The results of modelling the kinetic data revealed that soil pH primarily affected the slow desorption sites, and therefore variation in soil pH could induce a long-term effect on As[V] adsorption - desorption in soils (Mehmood et al. 2009).
2.5.2 Soil Organic Matter and Arsenic Phytoavailability
Organic matter has a diverse chemical nature and generally consists of numerous organic complexes having variable composition, structure and molecular weights (Shahid et al. 2012a). Soil OM consists of suspended and dissolved particles. Dissolved OM consists of non-humic (80%) and humic (20%) compounds (Rosas-Castor et al. 2014). These different forms of OM have variable effects on the geochemical behaviour of As in soils. Organic fractions having large molecular weights are capable to adsorb relatively more As whereas lighter fractions are usually more soluble and tend to dissolve metals either by displacing or chelating them (Moreno-Jiménez et al. 2012). Organic matter plays an important role in controlling the mechanisms of As binding in soils by forming ternary As complexes (Dousova et al. 2015).
Relationship between OM and As adsorption-desorption in soils is complex and is governed by various factors such as the Fe, Al and Mn concentrations present inOM, ratio of soluble and stable organic carbon (Gräfe and Sparks. 2006). Soil OM influences As geochemical behaviour in soils by changing its chemical speciation via modification in soil conditions. Organic matter can cause reduction of As[V] to As[III] as well as oxidation of As[III] to As[V] (Redman et al. 2002) depending on soil conditions. Soils having high level of OM induce reduced condition (Ryu et al. 2010) and alter redox potential in soils via proliferation of microorganisms. Redman et al. (2002) reported As[V] reduction into As[III] due to high soil OM. Several previous studies reported soil OM-mediated variation in As geochemical behaviour due to changes in soil properties (Rowland et al. 2006). Organic matter can change soil pH, microbial activity, redox potential, nutrient composition, and the available adsorption sites in soils (Wang and Mulligan 2006), which in turn affect geochemical behaviour of As.
Depending on the environmental conditions, soil OM exists in solid and soluble (dissolved) forms. In dissolved form, humic acid (HA) forms complexes with As and other cations. When present in solid phase, HA adsorbs As by binding it to functional groups. Certain studies supported OM-mediated decrease in As phytoavailability (Janoš et al. 2010) while others revealed an increase in As phytoavailability (McCauley et al. 2009). Soil OM-mediated increase in As mobility and phytoavailability can be due to the competition between dissolved OM and As for adsorption sites on mineral surface (Redman et al. 2002). This competition decreases As adsorption and enhances its mobility, especially under acidic conditions (Wang and Muligan 2006). Similarly, dissolved OM forms stable complexes with minerals block As desorption onto quartz, alumina, kaolinite or iron oxides.
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The complexation and/or adsorption of As by soil OM vary with the chemical speciation of As (As[III] vs As[V)]). The As[III] and As[V] have maximum adsorption onto HAs at pH 8.0 and 5.5, respectively (Ko et al. 2004). Similarly, solid phase HA has more affinity for As[V] than As[III]), with amine (-NH2) group possibly responsible for As binding and retention (Thanabalsingam and Pickering 1986). Despite complexation and binding of As by soil OM, total As in soil is poorly correlated with soil OM compared to Al, Fe, or P (Khalid et al. 2016), signifying limited role of OM towards As retention in soils.
2.5.3 Soil Microbes and Arsenic Phytoavailability
Microorganisms present in rhizosphere soil alter As solubility and chemical speciation thereby affecting As phytoavailability and mobility (Rahman et al. 2014). Microbes can either solubilize As, thereby enhancing its phytoavailability, or immobilize it thereby decreasing its phytoavailability. These bio-transformations of As play an important role in biogeochemical cycles of As and are exploited in bioremediation of As contaminated soils (Lloyd and Lovley 2001; Khalid et al. 2017b). Soil microbes increase As availability and solubilize As bearing minerals through different processes (Sheng and Xia 2006). Microorganisms play an important role in the redox transformations of As by reducing As[V]) to As[III] and vice versa (Gorny et al. 2015). A large number of As[V]) reducing and As[III] oxidizing prokaryotes have been found and isolated from the As contaminated environments (Ghosh et al. 2015) capable of facilitating the As transformations in agricultural soils (Corsini et al. 2010).
Incubation of the arbuscular mycorrhizal fungal strain for 10 weeks increased As phytoavailability for Zea mays plants (Bai et al. 2008). Similarly, Rahman et al. (2014) reported increased As uptake by Zea mays roots due to microbial interactions. The use of Acaulospora morrowiae and Glomus mosseae as arbuscular mycorrhizal affected the root As efflux and decreased the phytoavailability of As to the corn plants (Hua et al. 2014). Soil microorganisms may cause indirect As release in soil solution by the reductive dissolution of soil iron oxides. Several reports highlighted the ability of certain bacteria to release As by the catalytic reduction of ferric oxides (Bolan et al. 2014). On plant roots, iron plaque is formed by the abiotic oxidation and iron oxidizing bacteria (Weiss et al. 2003) which serves as ideal substrate for the iron reducing bacteria due to amorphous and crystalline structure (Hansel et al. 2001).
Microbes could transform As and other metals via methylation and demethylation; conversions of organic to inorganic forms and vice versa. Microbes obtain energy through the oxidation of As (Santini et al. 2000). Microbes cause oxidation of As[III] to As[V] which is precipitated through Fe ions (Chandraprabha and Natarajan 2011). Bacterial species such as Alcaligenes faecal, Geobacillus bacillus and Agrobacterium tumefaciens are capable of causing enzymatic oxidation of As[III]) into As[V] by synthesizing the arsenite oxidase (Yamamura and Amachi 2014). During oxidation of As[III] to As[V], bacteria obtain an electron for their metabolism. This As[III] transferring mechanism into less toxic form As[V] is regarded as a process of detoxification (Ghosh et al. 2015). In contrast, metals get reduced when microbes use As as a terminal electron acceptor for anaerobic respiration. Microbial-induced As[V] reduction in the uncontaminated soils is well-known (Xu et al. 2016) and also is main source of higher concentration of As in naturally contaminated waters (Lockwood et al. 2014). Microbial-induced As[V] reduction occurs through intracellular mechanism of detoxification (Ghosh et al. 2015) or dissimilatory reduction (Zobrist et al. 2000). Arsenate reducing microorganisms such as Geospirillum barnesi, Desulfutomaculum auripigmentu, Geospirillum arsenophilus, Bacillus arsenic oselenatis and Crysiogenes arsenates use As[V] as the terminal electron acceptor in their respiratory process (Vaxevanidou et al. 2012). The As[III] is more mobile and soluble compared to the As[V], therefore reduction of As[III] may result in the subsequent leaching (Bolan et al. 2014).
Transfer of toxic inorganic form of As to less toxic organo-arsenical form is termed as methylation that occurs through soil microbial activity. Microbial-mediated methylation of As plays key role in its biogeochemical cycle; As can be converted to gaseous arsines by biomethylation (Gao and Burau 1997). Methylation of inorganic As to organic As is considered a detoxification
39
mechanism adopted by some of the microorganisms (Zhang et al. 2015). Some microorganisms can transform As species into methylated forms such as mono-methylarsonic acid, di-methylarsinic acid and tri-methylarsine oxide (Gadd 1993). There is indication of As methylation in paddy soil systems where microorganisms transformed inorganic As into organic forms (Takamatsu et al.1982).
2.6 SOIL-PLANT TRANSFER OF ARSENIC
2.6.1 Molecular Understanding of Arsenic Absorption by Plants
Arsenic does not have any known essential role in metabolic and biochemical processes of plants and other living organisms. However, As is reported to be found naturally in plants. Although very rare, some researchers even revealed positive effect of As to plants at low exposure level (Gulz et al. 2005). Accumulation of As in plants generally remains <1 mg kg−1 (Adriano 2001). Arsenic content <0.1% of dry matter in Vicia faba, Solanum nigrum and Agrostis capillaris has been reported by Austruy et al. (2013) when these vegetables were grown on As polluted soils. Plants absorb As primarily through roots (Neidhardt et al. 2015), which can be via passive (does not require energy) or active (requires energy) pathways. Uptake of As by plants varies with its concentration and chemical speciation in soils, and is better related to exchangeable and bioavailable As level in soil (Martínez-Sánchez et al. 2011). Plants mainly absorb As as inorganic forms via transporter proteins, which is probably controlled by the concentration gradient of As between growth media and plant cells (Neidhardt et al. 2015). Among inorganic forms, As[V] is preferably absorbed by plants (Khalid et al. 2016).
No As-specific carrier for absorption by plants has been reported till date. The As[V] being a chemical analogue of P, plants compartment As[V] into plant cells generally through P cell channels using Pi (inorganic phosphorous) carrier system (Niazi et al. 2016). Physiological studies revealed that P and As[V] use the same absorption and transport pathway in plants (Meharg and Hartley-Whitaker 2002; Ghosh et al. 2015). By using synchrotron X-ray microprobe, Lei et al. (2012) revealed that P and As[V]) were co-transported via P channels. This co-transportation of As[V] and P can be enhanced inside plants by As[V] exposure or P deficiency. This is because of the fact that Pi transport system has much stronger affinity for P than that for As[V]. Hyper-tolerance in the plants against As[V] generally relies on reduced As[V] absorption by suppressing Pi uptake system.
Numerous P carrier proteins have been identified in plants (Bucher 2007). The P transporters which govern that of As[V] by plants consist of proteins called P transporter proteins (PHT). Nine PHT1-9 involved in the absorption and uptake of As have been identified in Mousear Cress. Although, it is reported that more than one PHTs are involved in As[V], but combinations of PHTs involved in As absorption and uptake of As still not well-known. The As[III] form is primarily absorbed and transferred within plants through members of nodulin-26-like intrinsic proteins (NIPs) (Khalid et al. 2016). Contrary to PHT proteins, NIP channels are bi-directional. Therefore, As[III] can be transported in both directions depending on the concentration gradient between plant cells and growth medium. Recently, Xu et al. (2015) reported that under As[III] stress conditions, Arabidopsis NIP3;1 plays key role towards As homeostasis. Compared to wild-type plants, the Arabidopsis NIP3;1 mutants acquired decreased As accumulation in plant aerial parts with increased tolerance. Similarly, NIP3;1 NIP1;1 double mutant exhibited enhanced roots and shoots growth and strong As[III] tolerance under high As[III] levels.
2.6.2 Arsenic and Complementary Cations
Adsorption of As and other metals is a competitive process between ions sorbed onto soil surface and those in solution. The competing cations include Mg, Ca, Fe, Na and K those could influence As sorption onto soil components and thereby affect its biogeochemical behaviour in soil-plant system These competing cations are held onto soil surfaces either by electrostatic attraction or through specific sorption. Both these processes play a key role in determining metal
40
compartmentation between solid and liquid phases. For example, it is shown that an increase in Mg2+ and Ca2+ concentration in natural and synthetic soil solution decreased As[V] affinity towards the mineral surfaces (Kanematsu et al. 2013). Plants are well-known to alter the physicochemical equilibrium in rhizosphere soil through the absorption of major anions and cations (Hinsinger et al. 2009). In rhizosphere soil, plant-induced decrease in Ca concentration in soil solution was correlated to increased level of P which competes with As[V] (Devau et al. 2010). Potassium and Ca both are reported to enhance As uptake by Poecilotheria vittata (Lombi et al. 2002). Calcium can enhance As uptake and accumulation in fronds of Poecilotheria vittata by increasing As translocation from roots to fronds (Tu and Ma 2002). A positive correlation between As and K accumulation in Poecilotheria vittata has been reported by Komar (1999). Similarly, presence of P in growth medium is used for As tolerance to non As hyperaccumulating plants by decreasing uptake of As[V] due to suppression by P uptake system (Tu and Ma 2003).
2.6.3 Arsenic Sequestration into Plant Roots
Plant ability to absorb metalloids is generally represented by the transfer factors from soil to plant (Xiong et al. 2014) and depends on as concentration and speciation in soil, type of soil, plant age, presence of the transporters/chelating agents in roots and type of plant species. Plant roots and tubers can accumulate As in larger quantity, which varies with the type of plant species. Arsenic transport to plant roots from soil depends on level of oxygen in rhizosphere soil. Soil-plant transfer of As greatly varies with the chemical form of As (Castillo-Michel et al. 2011). The As[III] has high soil-plant transfer ability under anaerobic condition whereas that of As[V] under aerobic condition (Oliveira et al. 2014). It is reported that soil-plant transfer of As is controlled by a complex mechanism of several factors such as soil pH, redox conditions, root exudates, root membrane activity, detoxification mechanisms, release of oxidants/redactors and phytosiderophores (Greger 2005). In majority of plant species, plant roots sequester major portion of As and only a small amount is translocated from roots towards shoots (Austruy et al. 2013; Rehman et al. 2016). Arsenic sequestration in plant root tissues can be as high as 200 times compared to that in shoots (Smith et al. 2008). This shows that plants roots are capable to accumulate most of the absorbed As in plants. The enhanced uptake and sequestration of As into plant root tissues have been reported in Oryza sativa (Rahman et al. 2011), Triticum aestivum (Shi et al. 2015), Helianthus annuus, Zea mays, (Neidhardt et al. 2015), Hydrilla verticillata (Xue et al. 2011) Vicia faba (Ye et al. 2010), Asplenium platyneuron (Niazi et al. 2016) and Solanum lycopersicum (Cobb et al. 2000).
2.6.4 Arsenic Translocation to Plant Shoots
Translocation ability of As is an important factor of plant as it represents the phytoaccumulation and remediation potential of a particular plant species (Vithanage et al. 2012). The root-shoot transfer factor is generally used to indicate potential of a plant species towards phytoremediation (Shahid et al. 2012b). Plants having root-shoot transfer factor values < 1 point out partial As transfer towards aerial tissues. Generally, As is considered less mobile regarding its transfer from plant roots to aerial parts (Vithanage et al. 2012). There exists great variation in translocation capacity of different plant species. Transfer of As ions towards plant shoots from roots is also governed by root pressure and leaf transpiration (Ghosh and Singh 2005). The As[III] mainly accumulates in xylem sap of Oryza sativa, Poecilotheria vittata, Solanum lycopersicum and Cucumis sativus, while As[V] predominated in xylem sap of Brassica juncea, Holcus lanatus, Ricinus communis and Triticum aestivum exposed to either As(V/III) (Ye et al. 2010). Hyperaccumulator plant species accumulate major portion of absorbed As in aerial parts of plants such as in xylem tissues, epidermal cells and mesophyll cells (Bondada et al. 2007; Niazi et al. 2016).
Translocation of As from roots to shoots varies greatly with its chemical form. Generally inorganic As species have higher translocation potential than organic As species. Abedin and Meharg (2002) showed that root-shoot transfer was several folds higher for As[III] or As[V] compared to MMA and DMA. Smith et al. (2008) revealed that aerial As was mainly in As[III] form
41
(57-78%) compared to As[V] (16 and 27%) and organic species in Oryza sativa. However, some researchers also reported contrasting results. Raab et al. (2007) showed that root to shoot translocation of DMA(V) was 3-folds higher than that of As[V] for 46 Zea mays species grown under hydroponics conditions. Some studies reported that Oryza sativa grains can accumulate up to 89% of DMA (Ackerman et al. 2005).
Arsenic is reported to interchange between As[V] and As[III] inside the plant body which affects its accumulation in different plant parts (Smith et al. 2009). Some plants are reported to reduce As[V] to As[III] inside the plant, and thus As translocation from root to shoot occurs as As[III] (Raab et al. 2005; Li et al. 2016). Mostly glutathione (GSH) is involved in reducing As[V] to As[III] inside the plants (Caille et al. 2004). It was suggested that reduction of As[V] to As[III] takes place in root cells before its loading into xylem and translocation towards aerial tissues. Generally, As[III] predominates (80%) in xylem sap of different plant species including Oryza sativa, Solanum lycopersicum and Poecilotheria vittata. However, this transformation of As[V] to As[III] is not always true because in some plants both forms of As[III] and As[V] are present in the xylem suggesting that both forms are loaded into xylem. Pickering et al. (2000) reported accumulation of As[V] at higher levels in comparison to As[III] in xylem sap of other plant species. Usually different transporters are involved in the transport of As from roots to shoots. Translocation of As towards aerial tissues in Oryza sativa was mainly arbitrated by silicon transporter Lsi2 (Ma et al. 2008).
2.6.5 Arsenic Accumulation in Edible Plant Parts
Plants absorb As mainly via root system from contaminated soils, and can accumulate in various parts of plants (Flora 2011). Arsenic is also reported to accumulate considerably in edible parts of plants. Arsenic translocation and accumulation in plant parts those are used as feed for human or animals which may induce serious health hazards. Therefore, now-a-days people are becoming increasingly concerned with respect to food safety owing to environmental pollution by As. Human exposure to high levels of As via contaminated foods induces noxious effects to different organs such as liver, kidney, and lungs, which may finally causes cancer. Therefore, As accumulation in edible parts and its possible consumption by living organisms have received significant consideration in the last decade.
Absorption and accumulation of As in edible parts of crop plants have been examined extensively to evaluate and to decrease the risk of As entrance into food chain. Studies recommend that majority of the crops do not translocate As to edible plant parts with aerial tissues generally low in As (<2 mg kg−1). However, some crops are capable to accumulate high levels of As in edible plant parts (5–40 mg As kg−1) (Gulz et al. 2005). Tripathi et al. (2012) reported that high amount of As (up to 149 mg kg−1) was accumulated in Oryza sativa straw. Generally leafy vegetables can absorb and translocate higher As levels into aerial edible parts compared to other grain crops/vegetables when cultivated under similar environmental conditions (Farid et al. 2003). Huang et al. (2006) reported the As accumulation trend in edible parts of crops in decreasing order as Brassica nigra > Spinacia oleracea > Lactuca sativa > Papuana uninodis > Brassica rapa > Vigna unguiculata > Brassica oleracea > Solanum melongena. Similarly, Fuente et al. (2010) reported As accumulation pattern in different crops in the decreasing order of Cichorium endivia ~ Brassica oleracea > Hordeum vulgare ~ Beta vulgaris ~ Triticum aestivum ~ Allium ampeloprasum > Capsicum anuum. High levels of As are also reported in Lactuca sativa, Solanum melongena and Allium cepa (Zhao et al. 2009).
2.7 TOXIC EFFECTS OF ARSENIC ON PLANTS
Arsenic toxicity to plants is a well-known phenomenon when its concentration in plant tissues increases a threshold level (Annex. Table 1). Plant roots are the first place which is exposed to As where this metalloid induces various noxious effects such as inhibition of root proliferation and extension. Translocation of As to shoots can severely impede plant growth. At sufficiently higher concentrations, As hampers critical metabolic and biochemical processes
42
which can result to even plant death. Majority of the plant species possess mechanisms for retaining much of their As burden in roots. Arsenic is reported to interfere with the numerous biochemical pathways such as impeded nutrient absorption, disruption of plant water status, replacement of essential ions from adenosine triphosphate (ATP), interaction with the functional groups of enzymes and negative effect on photosynthetic metabolism.
2.7.1 Arsenic toxicity to Plant Growth
Exposure of plants to high As results in physiological disorders and growth inhibition and finally may cause their death. Arsenic adversely affects seed germination and early seedlings growth (Abedin and Meharg 2002). There are increasing evidences that high concentration of As negatively affects various biochemical and physiological processes of plants leading to decreased economical yield (Khalid et al. 2016). It has been identified that As causes chlorophyll degradation, growth inhibition, nutrient depletion, photosynthesis activity diminution and membrane disintegration (Garg and Singla 2011; Finnegan and Chen 2012; Khalid et al. 2016). Arsenic also affects membrane system of chloroplasts, chlorophyll fluorescence and photosynthetic pigments, thereby decreasing photosynthetic activity (Ahsan et al. 2010; Garg and Singla 2011). Further, As has been shown to change the nutrient balance and their assimilation, protein metabolism and oxidative phosphorylation (Singh et al. 2009). For Agrostis capillaries and Vicia faba, As accumulation resulted in low root biomass especially in the Vicia faba which is highly sensitive to As toxicity (Duquesnoy et al. 2010).
Arsenic toxicity in plants causes various changes in plant tissues/cells such as wilting, necrosis, chlorosis, decrease in root length, changes on the chloroplast structure and modifications in the root cortical and epidermal cells (Campos et al. 2014). Accumulation of As in plants stimulates a series of events that can hamper plant growth, interrupt photosynthesis and respiration. Arsenic is also reported to mediate secondary metabolism which impedes plant growth and causes decrease in plant yield (Flora 2011). Arsenate toxicity could also decrease the chlorophyll biosynthesis in plants by interfering with photosynthetic apparatus. Decrease in pigment contents, root length and plant growth in response to As toxicity have been reported by various studies in plants (Finnegan and Chen 2012; Khalid et al. 2016).
2.7.2 Arsenic Genotoxicity
Arsenic is well-known to induce genetic damage in plants and animals. Arsenic causes induction of chromatid/chromosome breaks or exchanges (Patra et al. 2004; Patlolla et al. 2012), DNA and oxidative base damage (Snow et al. 2005), formation of apyrimidinic/apurinic sites (Yamanaka et al. 1997; Yi et al. 2007; Faita et al. 2013), DNA protein cross-links (Mouron et al. 2005), chromosomal aberrations, sister chromatid exchange, micronuclei formation, aneuploidy and deletion mutations. Although, numerous processes have been suggested to elucidate the cause of As-induced genetic damage, but still these mechanisms are not established. Studies have revealed that As-induced damage to DNA is not via direct interaction of As and DNA. Arsenic is generally reported a poor mutagen, because it is not bound covalently to DNA structures, thus does not induce any point mutations. Nevertheless, regardless of its poor mutation capacity, As is known to influence the mutagenicity of other carcinogens.
The primary phenomenon behind As-induced genotoxicity is via oxidative stress, i.e. over-production of ROS. It has been hypothesized that indirect toxic effects of As to plants genetic makeup via ROS production are more intense and quick compared to direct toxicity (Flora 2011; Farooq et al. 2016). Reactive oxygen species are known to initiate As genotoxicity, and are associated with the early steps of As-induced genotoxicity to plants (Shahid et al. 2015b). Over-production of ROS can cause DNA strand breaks, DNA adducts, chromosomal aberrations and cross-links (Shahid et al. 2014a). Arsenic can cause DNA base modification via ROS production, e.g., As can induce DNA nucleobase modifications by forming 8-oxoguanine (8-OHdG) adducts in DNA (Breton et al. 2007). Arsenic toxicity in plants also causes DNA strand breaks even at low exposure level. This DNA single-strand break caused by As can be during base excision repair process or by the action of ROS on the DNA bases (Gupta et al. 2011). Arsenic-induced
43
chromatid and chromosomal aberrations have also been reported in plants by using sister chromatid assay, micronuclei assay and chromosomal aberrations assay.
Arsenic-induced disturbances in essential biochemical processes results in chromosomal aberrations which impede DNA ligase (Patra and Sharma 2000). Arsenic can restrain multiple biochemical pathways linked with DNA methylation and DNA repair system (Sinha et al. 2013). Arsenic-induced over-production of ROS inhibits DNA repair mechanism by impeding base and nucleotide excision repair processes (Gupta et al. 2011). Arsenic can alter methylation state of DNA by modifying activities and amount of DNA methyl-transferase enzyme. Arsenic-induced alteration in DNA methylation process can occur directly via interaction between thiol enzyme and As or indirectly through ROS enzyme interaction. Moreover, As-induced ROS can also deplete S-adenosyl-methionine pool(s) which is necessary for methylation (Franco et al. 2008).
2.7.3 Arsenic-Induced Oxidative Stress
Inside the plant cells, As toxicity induces oxidative stress (Fig. 2.3, Table 2.1). The term ‘‘oxidative stress’’ refers to the enhanced production of ROS in cells due to toxicity of organic and/or inorganic pollutants (Zhao et al. 2009; Panda et al. 2010). The ROS are generally unstable, chemically very reactive and short lived molecules. These species contain unpaired electrons in their valence shell. These ROS include hydrogen peroxide (H2O2), singlet oxygen (½O2), superoxide anions (O2
• −), hydroxyl (HO•), alkoxyl (RO•), peroxyl (RO2•) radicals, and
organic hydroperoxide (ROOH). An enhanced production of ROS is generally the result of heavy metal toxicity including As in plant tissues (Shahid et al. 2014b). The ROS are naturally produced inside plants as by-products of normal aerobic metabolism in a number of plant organelles such as mitochondria, chloroplasts and peroxisomes (Singh and Upadhyay 2015). The ROS play important roles in plant metabolism which include: growth regulation, initiation of defence metabolism under stress, programmed cell death via signal transduction, fruit ripening and senescence, regulation of stomatal conductance, and alleviation of seed dormancy (Pourrut et al. 2011). However, the steady-state level of ROS under normal conditions is governed by the interactions between mechanisms responsible for ROS-production and ROS-scavenging.
When plants are exposed to heavy metals, including As, production of ROS increases that induce an imbalance between ROS generation and ROS scavenging (Sharma 2012). Over-production of ROS results in a number of physiological and biochemical disorders due to ROS interaction with proteins, DNA and lipids, enzyme inactivation and membrane leakage which can cause irreparable metabolic dysfunction and compromise the cell viability, thereby leading to cell death (Flora 2011; Shahid et al. 2014a; Farooq et al. 2016). Interaction of ROS with various biomolecules (such as RNA, DNA, lipids) and the consequential toxicity vary with the nature of ROS and the target molecules. The ROS-induced degradation of most of cellular components is generally permanent, only few biomolecules can be restored, such as DNA, cysteine and methionine residues in proteins (Pourrut et al. 2011). The proteins involved in generation and scavenging of ROS generally differ depending on the nature of As species in the plants body (Cuypers et al. 2012).
Arsenic is reported to cause oxidative stress in certain species of plants ensuring a range of changes in plant cells including re-adjustment of metabolic and transport processes (Talukdar 2013). Yamanaka and colleagues first reported As-induced ROS production in plants (Yamanaka et al. 1989). Arsenic can induce over-production of ROS in various ways; directly via the Fenton and Haber–Weiss reaction, and indirectly by decreasing the activities of various antioxidant enzymes (Flora 2011). Several previous reports revealed that As exposure causes over-production of ROS through conversion of As[V] to As[III] mechanism that takes place readily in plants (Gupta et al. 2013).
44
Figure 2.3: Vacuolar compartmentation of arsenic in plant cells.
Arsenic also causes ROS production during oxidation of As[III] to As[V] during the formation of intermediary arsine forms such as dimethylarsinic radicals [(CH3)2As•] and dimethyl arsinic peroxyl [(CH3)2AsOO•] (Yamanaka and Kada 1994). The ROS are generated due to the electron leakage during inhibition of the key enzyme systems, As[III]-As[V] reduction which results in number of damaging effects including GSH depletion, membrane leakage and decreased photosynthetic activity (Sharma 2012). Furthermore, As stress in plant cells enhances the consumption of oxygen, thereby increasing ROS production. Arsenic can bind directly to histidyl-, thioyl- and carboxyl-groups, which is responsible for As-induced ROS production.
Arsenic-induced ROS mainly include •OH, H2O2, O2•−, 1/2O2, and peroxyl radicals. The
class of radical species derived from oxygen is considered most important owing to a distinctive electronic configuration of O2
•− after accepting one electron (Miller et al. 1990). Superoxide produced via metabolic processes is considered “primary” ROS, which can generate “secondary” ROS after interaction with enzymes and molecules (Sharma 2012). This enhanced production of ROS causes changes in the redox status of cells, which causes activation of antioxidant mechanisms. Over-production of ROS not only causes widespread damage to various molecules but also functions as signalling molecules. Arsenic-induced increase in ROS has been reported in many plant species such as Arabidopsis thaliana, Phaseolus aureus, Nasturtium officinale, Luffa acutangula, Myracrodruon urundeuva and Oryza sativa (Gupta et al. 2013; Malik et al. 2012; Namdjoyana and Kermanian 2013; Singh et al. 2013; Gomes et al. 2014; Dixit et al. 2015).
2.8 ARSENIC DETOXIFICATION IN PLANTS
The ROS are produced naturally in plants during different metabolic processes; however, their production is enhanced under stress conditions. There exists numerous plant species, especially hyperaccumulators of metalloids, which can survive and grow well under high ROS
PHT NIPsAs(V) As(III)
GSHGSSG
Arsenate Reductase
H2O2•OH + OH− O2O2
-
MMA, DMA
Oxidation
As(III)-PC
PC + As (III)
Vacuolar compartmentation
Plasma
membrane
Reduction
Oxidative stress
As(V) As(III)
45
production situation. In order to survive and cope with metal stress, plants are equipped with the composite mechanisms of defence against ROS (Farooq et al. 2016). Plant defence systems operate in conjugation with each other to detoxify ROS over-production. Nonetheless, which particular plant defence mechanism will operate under specific conditions depends on various factors: type and level of metal, duration of exposure, type of plant species etc. (Baxter et al. 2014). These defence mechanisms identify and mediate tress signals to adjust their metabolic processes according to stress conditions (Sharma 2012). Different strategies developed by plants against As toxicity include chelation of As with phytochelatins (PCs) and GSH reduction of As[V] to As[III] and compartmentation of As in vacuoles (Flora 2011).
2.8.1 Vacuolar Compartmentalization of Arsenic in Plants
Plants can avoid As toxicity by sequestering and storing it into vacuoles (Fig. 2.3) (Zhao et al. 2009; Flora 2011). Certain plant species especially the hyperaccumulator plants have ability to sequester and/or bind metals to molecules or plant structures that limits negative effects. Heavy metals are detoxified in aerial tissues of hyperaccumulator plants due to their binding with PCs and sequestration into vacuoles (Zhao et al. 2009). Vacuolar transporters partially accomplish this task, causing the partitioning metals into vacuole (Shahid et al. 2014a). Cell vacuoles are considered an end point for almost all the toxic elements. However, researchers have proposed several mechanisms of metal complexation and transportation into vacuole.
Table 2.1: Arsenic-induced enhanced production of ROS in various plant species grown in
hydroponic culture
ROS Plant species As exposure level Duration References
H2O2 Arabidopsis thaliana
25, 50 µM Na2HAsO4 1, 5 days Gupta et al. 2013
H2O2 Phaseolus aureus 2.5, 5, 10 µM Na2HAsO4 10 days Malik et al. 2012
H2O2 Nasturtium officinale
100 µM Na2HAsO4 7 days Namdjoyana and Kermanian 2013
H2O2 Luffa acutangula 5, 50 µM Na2HAsO4 7 days Singh et al. 2013
H2O2 Myracrodruon urundeuva
10, 50,100 mg kg−1Na2HAsO4
120 days Gomes et al. 2014
H2O2 Oryza sativa 25 µM NaAsO2, 50µM
Na2HAsO4 7 days Dixit et al. 2015
H2O2 Pusa Jai Kisan 5, 25 µM Na3AsO4 4 days Khan et al. 2009
O2•−, H2O2 Oryza sativa 10, 25 µM NaAsO2 10 days Tripathi et al. 2013
H2O2 Pteris vittata 133 or 267 µM Na2HAsO4 1, 5, 10 days Singh et al. 2006
H2O2 Pteris ensiformis 133 or 267 µM Na2HAsO4 1, 5, 10 days Singh et al. 2006
46
H2O2 Oryza sativa 100 µM Na2HAsO4 2 days Nath et al. 2014
H2O2 Cicer arietinum 30, 60 mgkg−1NaAsO2φ 48 days Gunes et al. 2009
H2O2 Oryza sativa 0.1, 0.5 or 2.5 mg/l NaAsO2
14 days Ren et al. 2014
O2•−, H2O2
Solanum lycopersicum
53.6 µmol/L NaAsO2 4, 8, 12, 16, 20, 24, 36, 48, 72 h
Singh and Upadhyay 2015
O2•−, H2O2 Pistia stratiotes
3, 7, 10, 13, 16, 20 µM Na2HAsO4
7 days Farnese et al. 2014
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide.
Numerous transporter families mediate As and other heavy metals transportation towards vacuoles in plants (Flora 2011). These transporter families include cation diffusion facilitators (CDF), natural resistance-associated macrophage (NRAMP), ZIP (ZRT, IRT-like protein), heavy metal ATPases (HMAs), ATP-binding cassettes (ABCs) and cation exchangers (CAXs) (Zhao et al. 2009; Shahid et al. 2014a). Among these, ABC, NRAMP and CDF have been recognized as major heavy metal tolerance transporters (Zhao et al. 2009; Chaffai et al. 2011). These transporter proteins chelate As generally through thiols group (Finnegan et al. 2012) before transportation into vacuoles (Sharma 2012). It is supposed that detoxification and the vacuolar compartmentalization of As are vital in all the plant organs. Some authors have reported that As complexes with transporter proteins may dissociate inside vacuoles when pH is around 8, and ligand can again make complexes and transport As into vacuoles. The ATP-binding cassette super-family generally facilitates As complexation by thiol groups and then mediate its compartmentation towards vacuole (Verbruggen et al. 2009).
2.8.2 Arsenic Chelation by Phytochelatins
Phytochelatins (PCs) are heavy metals chelating cysteine rich polypeptides with the general structure (c-Glu-Cys) n-Gly (Cobbett 2000; Hawrylak-Nowak et al. 2014). Phytochelatins were first identified and isolated from cell cultures of higher plants after Cd stress (Grill et al. 1985; Cobbett 2000). Since their discovery, the PCs have been identified in numerous plant species. Thereafter, several physiological studies indicated the role of PCs in the homeostasis and detoxification of toxic metals as well as in the conservation of ionic homeostasis.
Phytochelatins are among the most important plant molecules involved in As and other metals detoxification in plants (Finnegan et al. 2012). In addition to plants, PCs are also found in the fungi and other organisms (Piechalak et al. 2002). Phytochelatin synthase (PC-synthase) uses GSH as a substrate for the production of PCs (Flora 2011). The PCs form complexes with As and other metals in the cytosol followed by their compartmentation into vacuoles (Fig. 2.3; (Meharg and Hartley-Whitaker 2002; Shahid et al. 2014a). In this way, PCs decrease the concentration of As in the cytosol and inhibit ROS production, and thereby defend plants against harmful effects of As. Synthesis of PCs under As stress conditions is possibly one of the most well studied process governing metal tolerance in plants. Important role of PCs in detoxification of As was reported for various plants species such as Cytisus striatus Silene vulgaris, Arabidopsis, Holcus lanatus and Rauvolfia serpentina (Bleeker et al. 2003; Zhao et al. 2009).
There is strong indication that PCs govern an important role in As detoxification in terrestrial plants. The complexation of As with PCs is via thiol-rich peptides (Sharma 2012). Moreover, the transportation and sequestration of As in vacuole occur via ATP binding cassette
47
transporters (Zhao et al. 2009; Finnegan et al. 2012). Nevertheless, pumping of PC–As complexes into vacuoles is not well clear so far. Wojas et al. (2009) reported that over expression of AtPCS1 or CePCS resulted in an increased As accumulation and tolerance in Nicotiana tobacum plants. Similarly, when PC synthesis was blocked by potent inhibitor l-buthionine-sulphoxime, it caused As hypersensitivity (Schat et al. 2002). Plants can accumulate both As[III] and As[V], former being more toxic to plants. Therefore, one of the mechanisms adopted by plants against As detoxification is As[V] reduction into As[III] followed by complexation of As[III] by thiol groups and As-PC complexes compartmentation in vacuoles (Fig. 2.3) (Zhao et al. 2009; Finnegan et al. 2012). It has been shown that wild type plants showed more resistance and tolerance to As toxicity than PC deficient mutants of Arabidopsis thaliana (Liu et al. 2010). Several researchers reported that over expression of PC-synthase increased As homeostasis in transgenic plants (Flora 2011). When the synthesis and over expression of PC were inhibited using protein inhibitors (L-buthionine-sulphoximine), plants showed high sensitivity to As (Schat et al. 2002).
Although over production of PCs in plants is well-known to enhance As accumulation and tolerance, but some studies debate that some plants, especially hyperaccumulators, could tolerate As without over production of PCs. For instance, As hyperaccumulator Pteris vittata used PC independent mechanism of As detoxification (Raab et al. 2004). They reported that only small proportion of As was absorbed by Pteris vittata that was complexed by PCs (Webb et al. 2003). These plants can reduce As[V] to As[III], which is transported and stored in vacuoles generally in non-complexed form (Lombi et al. 2002) possibly via tonoplast transporter PvACR3 (Indriolo et al. 2010). Similarly, Zhao et al. (2003) stated a limited role of PCs in the As[III] detoxification of Poecilotheria vittata because of low PC-SH to the As molar ratio of 0.09:0.03.
2.8.3 Arsenic Chelation by Glutathione
Glutathione is one of the most important low-molecular weight thiols. Glutathione (γ-glutamate cysteine-glycine) is a sulfur- containing tripeptide thiol. Glutathione plays a key role in the defence and tolerance against As-mediated oxidative damage by participating in a number of biochemical reactions such as reduction of peroxides, modulation of thiol-disulfide status and free radical scavenging (Foyer and Noctor 2005; Zhao et al. 2009). Two ATPM dependent enzymes (glutathione synthetase and γ-glutamyl cysteine synthetase) catalyse the synthesis of GSH. Glutathione can neutralize the As-induced over production of ROS through ascorbate-glutathione cycle (ASA-GSH cycle), which takes place in different sub-cellular compartments (Foyer and Noctor 2005). Glutathione plays an important function in plants through its thiolic residue of the cysteine (Cys) therefore, functioning as key controller of the redox homeostasis (Zhao et al. 2009; Srivastava et al. 2014). Glutathione-s-transferases (GST) have several functions like metals detoxification in primary and secondary metabolism in signalling and stress tolerance.
Glutathione is also linked to other defence pathways such as ASA-GSH cycle and PC synthesis. The As[III]and As[V] have high binding strength for thiols peptides such as GSH (Finnegan et al. 2012). It is reported that arsenate reductase (AR) mediates the reduction of As[III] to As[V]; the mechanism known for the detoxification of As in plants because GSH can bind As[III]. During reduction of As[V] to As[III], nicotinamide adenine dinucleotide phosphate (NADPH) gets oxidized through the reduction of oxidized glutathione (GSSG) by glutathione reductase. During this process, GSH serves as an electron donor for the AR (Ellis et al. 2006). Moreover, GSH can enhance As tolerance indirectly by binding ROS and detoxifying it by the process catalysed by the GST. Arsenic has been reported to activate GST in several plants (Sharma 2012). However, there exists less data regarding the effect of As on the GSH level in plants compared to other heavy metals. Singh et al. (2006) reported enhanced production of GSH in As hyperaccumulator Poecilotheria vittata mediates As[V] tolerance. Substantial increase in GSH pool after As exposure has been documented in numerous tolerant plants like Hydrilla verticillata (Flora 2011). In Oryza sativa, supplementation of GSH enhanced As tolerance and inhibition of the oxidative stress (Shri et al. 2009).
48
Enhanced production of GSH is not always true under As stress conditions. Some plants can tolerate As-mediated oxidative damage without producing GSH. Even some researchers reported depletion in GSH pool under As stress in plants. For example, Hartley-Whitaker et al. (2001) suggested that in Holcus lanatus rapid As[V] influx resulted in GSH depletion and PCs production. Similarly, decrease in GSH contents was witnessed in Trifolium pratense plant upon the exposure of 50 mg kg−1 As (Mascher et al. 2002).
2.8.4 Antioxidant Enzymes and Arsenic Toxicity
One of the most competent mechanisms that make plants tolerant to As-induced over-production of ROS is the presence of antioxidative defence system (Finnegan et al. 2012). Such defence system averts As-induced tissue dysfunction and cell injury and is mediated in plant issues by the activation of antioxidant enzymes. Arsenic enhances the generation of ROS (Talukdar 2013) resulting significant damage to plant metabolism. Inside plants, ROS are scavenged by a complex system of both non-enzymatic antioxidant (GSH, carotenoids and ascorbate (ASC) and enzymatic antioxidative system [guaiacol peroxidase (GPX), catalase (CAT), lipoxygenase (LOX), superoxide dismutase (SOD), ascorbate peroxidase (APX), and glutathione reductase (GR)] as given in Table 2.2 (Talukdar 2013). Arsenic-induced increase in activities of SOD, CAT, APX and GR has been reported in Triticum aestivum, Phaseolus aureus, Brassica juncea, Oryza sativa, Vicia faba and Zea mays (Liu et al. 2007; Malik et al. 2012; Namdjoyana and Kermanian 2013; Kanwar et al. 2015; Shri et al. 2009; Duquesnoy et al. 2010; Duquesnoy et al. 2010).
The antioxidant enzymes work in conjugation with each other to scavenge ROS. Among these enzymes, CAT which is generally present in peroxisomes and mitochondria decomposes H2O2 by an energy efficient mechanism (Shahid et al. 2014a). The CAT is a tetrameric heme having enzyme present in different cell organelles such as mitochondria, glyoxysome, cytosol, peroxisomes and root nodules. Exposure to As may enhance or decrease CAT activity in plants. Arsenic exposure enhances CAT activity by increasing H2O2 production. On the other hand, As-induced decrease in CAT activity can be due to As toxicity or direct binding of As to functional groups of the CAT. Arsenic-mediated increase in CAT activity has been described in Wedelia chinensis (Talukdar 2013). In contrast, As-induced decrease in the CAT activity has been described in the Taxithelium nepalense and Vicia faba (Singh et al. 2007). Activation and intensity of activation of the CAT also vary with plant type. For example, As tolerant Pteris vittata showed higher CAT activity than the As sensitive Nephrolepis exaltata and Ptries ensiformis (Shrivastava et al. 2005).
The superoxide dismutase is an important enzyme which plays central role for scavenging ROS. This enzyme dismutates two O2
• − radicals to O2 and H2O2, and thereby controls steady-state level of O2
•− in plant cells. Arsenic-induced activation of SOD could be due to an increase in O2
• − level or direct action on SOD. Increase in SOD activity can also be due to de-novo synthesis of enzyme protein (Yılmaz and Parlak 2011). The SOD is related with the metal co-factors; Mn-SOD enzyme is present in mitochondria, Fe-SOD is present in plastids and Cu/Zn-SOD is found in plastid, peroxisome, cytosol and root nodules. Increase in SOD activity as a result of As toxicity has been reported in Vicia faba, Zea mays and Trifolium repens (Mascher et al. 2002; Duquesnoy et al. 2010). Arsenic-induced high response of the Cu/Zn SOD has been demonstrated in Oryza sativa (Shri et al. 2009). Arsenic exposure to Zea mays roots showed up-regulation of Cu/Zn SOD which resulted in cellular homeostasis against As during the redox disturbance (Duquesnoy et al. 2010). The GPXlocated in cell walls or the cytoplasm is a member of the large peroxidase family. The GPXs detoxify H2O2 in plants (Foyer and Noctor 2005) as well as phospholipid hydroperoxides and lipids (Avery and Avery 2001). Increased activities of GPX on exposure to As have been shown in Myracrodruom urundeuva (Gomes et al. 2014). An increase in GPX activity has been demonstrated in Pteris vittata (Raj et al. 2011), Oryza sativa (Dixit et al. 2015) and Myracrodruom urundeuva (Gomes et al. 2014). The GST, formerly recognized as ligandins, include a family of prokaryotic and eukaryotic phase II metabolic isozymes. These enzymes catalyze the conjugation of reduced GSH to xenobiotic substrates during detoxification. The GST contains six functional classes in plants; DHARs, phi, theta, tau,
49
zeta and lambda (Ghelfi et al. 2011). Arsenic-induced activation of GST has been reported in Luffa acutangula and Pteris vittata (Paul and Shakya 2013; Singh et al. 2013).
Table 2.2: Arsenic-induced activation of antioxidant enzymes in different plant species grown in
hydroponic culture
Enzymes Plant species Asexposure level Duration Reference
APX, GPX, CAT, GSH Arabidopsis thaliana
50 µM Na2HAsO4 7 days Dixit et al. 2015
SOD, CAT, APX, GPX, AsA Phaseolus aureus 10, 50,100 mg kg−1 Na2HAsO4 120 days Gomes et al. 2014
SOD, POX, CAT, APX Nasturtium officinale
6.6, 13. 2, 26.4, 152.8 µM L- 1 Na2HAsO4.7H2O
3 days Gusman et al. 2013
SOD, CAT, POD Luffa acutangula 1, 2, 4, 8 mg/L Na3AsO4.12H2O
2 days Liu et al. 2007
SOD, CAT, GR, APO, ASC Myracrodruon urundeuva
2.5, 5, 10 µM Na2HAsO4.7H2O
10 days Malik et al. 2012
CAT, APX, SOD Oryza sativa 100 µM Na2HAsO4.7H2O 7 days Namdjoyan 2013
SOD, CAT, POD, APX, GR, MDHAR, DHAR
Pusa Jai Kisan 0.1, 0.3, and 0.5 mM Na2HAsO4.7H2O
60 days Kanwar et al. 2015
GR, SOD, APX, POD Oryza sativa 50, 100 µMNa2HAsO4 10 days Shri et al. 2009
SOD, CAT, POX Pteris vittata 134 and 668 µM As2O5 8 days Duquesnoy et al. 2010
SOD, CAT, POX Pteris ensiformis 135, 668 µM As2O5 8 days Duquesnoy et al. 2010
SOD, GSH Oryza sativa 5, 10, 50mg kg−1Na2HAsO4 φ 70 days Mascher et al. 2002
SOD, CAT, AsA, GSH Cicer arietinum 100 µM Na2HAsO4 2 days Nath et al. 2014
GR, GSH Oryza sativa 80, 160, 320, 640, 1280 µM AsNa2O4
1 day Goupil et al. 2009
SOD Solanum lycopersicum
5.0, 7.5, and 10 mg/ml As2O3 180 days Paul and Shakya 2013
SOD, CAT, APX, GST Pistia stratiotes 5, 50 µM Na2HAsO4. 7H2O 7 days Singh et al. 2013
GSH, APX, GR, GST Arabidopsis thaliana
5-50 mg kg−1Na2HAsO4. 7H2O 45 days Raj et al. 2011
SOD, CAT, APX Phaseolus aureus 25, 50, 100 μM As2O
20 days Shukla et al. 2015
50
APX, GSR, CAT Zantedeschia aethiopica
34.5 μg/L As 180 days Del-Toro-Sánchez et al. 2013
SOD; superoxide dismutase, APX; ascorbate peroxidase, GR; glutathione reductase, POD; Peroxidase, LOX; lipoxygenase, GPX; guaiacol peroxidase, CAT; catalase, AsA; ascorbic acid, GST; Glutathione S-transferase, AAO;
ascorbic acid. φ Grown in soil
Peroxidase is a big family of enzymes that normally catalyses the reduction of H2O2 into H2O. In plant cells, ascorbate acts as reductant for the reduction of H2O2 (Mehlhorn et al. 1990). The APX is a heme-containing protein present in membrane and plastid stroma. The APX and two molecules of ascorbate catalyse the reduction of H2O2 into H2O. During this process, two molecules of the monohydro-ascorbate are produced (Nactor and Foyer 1998). Up-regulation of the APX activity under As exposure has been stated in Wedelia chinensis (Talukdar 2013), Oryza sativa (shri et al. 2009), Zantedeschia aethiopica (Del-Toro-Sánchez et al. 2013) and Hordeum vulgare (Shukla et al. 2015). The GR is involved in maintaining GSH level in plant cells (Shahid et al. 2014a). Enhanced activity of GR has been reported in roots of Nephrolepis exaltata, Ptries ensiformis and Pteris vittata (Goupil et al. 2009; Kanwar et al. 2015). The GR mediates the enhanced necessity of GSH during As-mediated oxidative stress in plants (Shri et al. 2009).
2.8.5 Salicylic Acid and Arsenic Toxicity
Salicylic acid (SA) and its derivative acetyl-salicylic acid are hormone-like signalling substances that affect numerous developmental and physiological mechanisms. Studies reported a key role of SA towards regulation of plant growth, flowering, seed germination, glycolysis and fruit yield (Guo et al. 2009). Moreover, SA mediates specific reactions as a result of oxidative stress induced by As. Salicylic acid can enhance plant tolerance by reducing As-mediated oxidative damage in plants (Vlot et al. 2009). This shows that SA plays an important role in plant defence to As stress (Tao et al. 2013). The defensive role of SA mainly includes antioxidants activation, distribution of essential nutrients, gene regulation and membrane stabilization (Belkadhi et al. 2015). Arsenic exposure enhances SA contents in various plant species (Tao et al. 2013), but it is not still well-known that how SA is synthesized and regulated after As and other metals stress. It is proposed that SA may be synthesized by the plants via two distinct pathways: (i) by using l-phenylalanine in phenylpropanoid pathway and (ii) by using metabolite chorismate during compartmentalized enzymatic pathways.
Salicylic acid-induced tolerance of As also holds for exogenous applications of SA, which decreases As-induced production of ROS, TBARS, and lipid peroxidation, and mitigates the inhibition of antioxidant activities in plant cells (Panda and Patra 2007). Singh et al. (2015) reported that exogenous applications of SA decreased significantly As level in root and shoot tissues of Oryza sativa. Salicylic acid is known to decrease As translocation towards shoot tissues which can be due to SA-induced regulation of root-shoot As transporters. Salicylic acid can activate ABC transporters in the Glycine max (Eichhorn et al. 2006), which are involved in vacuolar compartmentation of As[III]-PC complexes (Briat et al. 2010). Consequently, it is proposed that SA plays a key role in sequestration of As-PC complexes in vacuoles of root cells, thereby, decreasing As translocation towards shoot tissues. Moreover, exogenous application of SA also helps increased PC synthesis in Zea mays root (Szalai et al. 2013).
2.9 HORMETIC EFFECTS OF ARSENIC TOXICITY
The hormetic responses in plants have received comparatively little consideration until now (Calabrese 2005). Wiedman and Apple (1972) studied plants hormesis for the first time. Since then different chemical doses leading to the bi-phasic responses have been documented including As and other heavy metals in plants. Generally, As is considered non-essential for living organisms, however, some studies have reported the beneficial effects of As to plants at very low
51
concentrations (Gulz et al. 2005). However, these studies reporting beneficial effects of As on plant growth are very rare. Recently, Rafique and Shahid (2016) reported positive effects of As on growth parameters of three vegetables, i.e. Vicia faba, Pisum sativum and Spinacia oleracea. This bi-phasic dose–response mechanism is mostly reported by an inverted U- or J- shaped curve. The shape and slope of curve vary with the process/parameter studied. For instance, a hormetic response curve with inverted U-shape is usually achieved for normal physiological parameters, such as plant growth. In addition to As, dose-response hormetic effect has frequently been reported on plant growth under low level exposure of toxic metals such as Pb, Cd, Al and Cr (Poschenrieder et al. 2013).
2.10 CONCLUSIONS AND PERSPECTIVES
Arsenic does not play any essential role in biochemical reaction of plants. Arsenic is released into soil both by natural and anthropogenic sources. Arsenic occurs in soil/solution in various chemical (especially arsenate and arsenite) forms which differ significantly with respect to their mobility/solubility, phytoavailability and toxicity. Chemical speciation of As [As(III) or As(V)] greatly affects its soil-plant transfer and phytoaccumulation. Till date, no As-specific transporter has been identified in plants. Arsenic enters plants mainly via root cells by phosphate and silicon transporters. Arsenic accumulates primarily in plant roots with limited translocation to shoot tissues. Inside plants, As inhibits seed germination, water uptake absorption and various metabolic processes. It also induces production of ROS and causes DNA damage and lipid peroxidation. Arsenic tolerant plant species sequester As in vacuoles as a detoxification mechanism. Plants tackle As toxicity via different defence mechanisms such as decreased As phyto-uptake or sequestration into the vacuole. Numerous antioxidants control As toxicity by scavenging As-mediated ROS. Other defense systems involve induction of glutathione, phytochelatins and salicylic acid.
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CHAPTER-3
BIOGEOCHEMICAL BEHAVIOUR OF LEAD IN SOIL-PLANT SYSTEM
3.1 INTRODUCTION
Lead (Pb, atomic number 82) is a naturally occurring metal belonging to group IV in the Periodic Table. This metal is bluish-grey, dense, ductile, poor electrical conductor and resistant to corrosion. Owing to various useful physicochemical characteristics, Pb is mined since antiquity and has a long history of usage in industrial processes (ATSDR 2007b; Paoliello and Capitani 2005). Lead has been used by ancient civilizations of Egyptians, Romans and Greeks for making water pipes and lining baths, and the plumber who mends and joins pipes (Gutiérrez et al. 2016). By the time of Augustus, Pb has been in use by the Romans to make brass (Gutiérrez et al. 2016). Lead is ranked as fifth metal behind zinc, copper, iron and aluminium in the industrial production of metals. Lead continues to be used extensively in numerous manufacturing processes, which has caused a noteworthy higher concentration of this metal in different compartments of environment. i.e. water, soil and air (Paoliello and Capitani 2005; ATSDR 2007b; Pourrut et al. 2011). The major sources of environmental contamination by Pb include mining, burning of fossil fuels, plumbing, weapons manufacturing, paints based on Pb, x-ray shields, ammunition, stained glass windows, containers for corrosive liquids and solder (ATSDR 2007b; Pourrut et al. 2011; Gutiérrez et al. 2016). High Pb contents can also build-up in soil with the application of sewage sludge as well as release of treated effluents (Murtaza et al. 2010).
The release of Pb into soil and water can be a source for exposure to living organisms. Lead is a non-mobile and less soluble compound due to formation of stable complexes with soil constituents and has a tendency to be stored mainly in soils. However, mobile fraction of Pb in soils can be absorbed by plants (Shahid 2010) and thereby contaminate the food chain. Absorption of Pb by plants is governed by numerous soil factors such as soil structure, texture, pH, root surface area, root exudates and concentration of Pb in soils (Shahid et al. 2012a). Lead mainly enters into plants through root cells by Ca permeable channels (Pourrut et al 2008). After absorbance by plant roots, Pb is not easily transferred to aerial shoots of most of the plants, but is sequestered mainly into root cells. It is reported that about 95% or more of Pb absorbed by most plants is accumulated in plant root cells except the plant is either chelate-assisted or Pb hyperaccumulator (Sharma and Dubey 2005; Saifullah et al. 2015). Lead does not accumulate primarily in edible parts of fruit and grain crops (Cucurbita maxima, Solanum lycopersicum, Zea mays, Fragaria ananassa, Malus domestica and Vicia faba). Higher concentrations of Pb in leafy vegetables are more likely to be found (e.g., Lactuca sativa) and on the surface of root vegetables like Daucus carota (Khalid et al. 2016).
Lead does not play any known biological role in metabolic and biochemical processes of living organisms. Generally, plants are equipped with an efficient defence system against Pb stress that limits its interactions with sensitive biological tissues and thus safeguards plants. However, presence of Pb inside plant cells can induce a whole range of harmful effects to plants. Accumulation of Pb at higher levels in plants can induce several toxic effects such as decreased seed germination, inhibition of chlorophyll synthesis and plant biomass (Pourrut et al. 2011). Inside plant cells, Pb can negatively affect enzymatic reactions, mineral nutrition, respiration, photosynthesis and numerous other physiological parameters (Sharma and Dubey 2005). Lead-induced ROS is considered an early and major cause of Pb toxicity to plants, which can disturb cell redox status, thereby inducing oxidative stress in plant cells (Pourrut et al. 2013).
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Lead toxicity and poisoning have been reported throughout the history of human civilizations. Lead is a hazardous chemical; it could not only pile-up in individual organisms but can also build-up in entire food chain. There is significant concern about human exposure to Pb due to its numerous noxious health hazards on hematopoietic system and neurological development in young children particularly (ATSDR 2007b). Human exposure to Pb occurs through dermal contact, ingestion and inhalation (ATSDR 2007b). Soil ingestion is considered a significant exposure path for Pb through outdoor activities in children that could cause ingestion of 50–200 mg soil/day on an average for a child (Oomen et al. 2003). Lead can induce disorder in a number of biochemical and metabolic processes taking place in human body. Lead poisoning can induce mental impairment in children of age < 15 (Wilhelm et al. 2010). Lead poisoning can seriously affect body organs such as bones, heart, kidneys and intestine as well as nervous and reproductive systems. Lead is well-known to interfere with the development of nervous system and thereby can cause toxicity to children. Symptoms of Pb poisoning include headache, confusion, abdominal pain, irritability and anaemia. Under sever Pb toxicity, coma, seizures, and even death may occur. Due to these and numerous other known hazardous effects, Pb is considered as a substance of high concern in the European regulations. Moreover, Pb ranks 2nd among hazardous substances listed by ATSDR (2007b).
3.2 GLOBAL USES OF LEAD
Lead, having a wide range of unique properties, has been widely used by human beings in different products and applications for centuries. Lead demand was higher in industrial age because of its high chemical stability in soil, air and water. High resistance and poor electrical conductivity of Pb are among the important characteristics because of having corrosive liquid such as the sulfuric acid. Lead in the refined form in mined ore has been produced since ancient times. Major producers of Pb are Australia, USA and China followed by Peru, Sweden, Mexico and Canada (USGS 2016). During urbanization and development process in China, polluting enterprises had been shifted to rural and suburban areas. In the world, Pb battery factories with a production of about 1/3rd of total Pb in the world are a substantial source of Pb pollution. Over time its consumption has been changed, as in paint pigments and gasoline additives in the final stages of phased out and also at present being used in the Pb acid batteries (Mao et al. 2008). Lead releases into the air in USA decreased from 221 thousand tons in 1970 to 160, 74, and 4-5 thousand tons, respectively in 1975, 1980 and 1990. This significant decrease shows the enormous impact that leaded gasoline had on Pb release into the environment.
Global inventory in late 1980s by Nriagu and Pacyna (1988) showed that about 356-857 ×106 kg Pb was emitted into the environment through smelting and mining activities annually. Lead is emitted into the atmosphere since the establishment of modern industries and medieval and in ancient times since metallurgy birth (Klaminder et al. 2008). Lead enters into the environment easily following atmospheric fallout emitted from the industrial matters that will affect development and growth of plants and also affects metabolic activity of micro-organisms in soil system even it destroys the ecosystem's balance completely. Recently, anthropogenic activities especially usage of Pb storage battery, manufacturing industry and vehicle exhaust have devoted to Pb contamination (Shahid et al. 2014c). Previous studies show that contribution of Pb into the environment due to long term anthropogenic activities has resulted in contamination of Northern Hemisphere. Lead is the most ductile metal. In addition to its wide-spread use in building material and piping, Pb and its oxide have been extensively used in paint, pewter and Pb glazes since Roman times. In 20thcentury, tetraethyl Pb was widely used to petrol as an anti-knock additive. Currently, Pb commercial uses still include building material (gutter material and sheeting), car batteries, cable sheathing, ammunition, Pb crystal glass, solders and radiation protection. The main Pb source is the sulfide ores although today just half of total annual production of Pb near 8 million tons is basically from recycled Pb. According to USGS (2016), Pb mine production has increased by 26% from 1,580x106 metric tons to 2,530x106 metric tons during 2006 to 2015, respectively (Fig. 3.1).
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Figure 3.1: Annual world refinery production of lead (USGS 2016).
3.3 LEVELS AND SOURCES OF LEAD IN SOILS
Lead is the 36thmost abundant component of the Earth crust (John 2011). Different authors/reports described different average concentration of Pb in the Earth crust and surface soil. For example, average concentration of Pb in soil is 13 mg kg-1 (Nriagu 1978), 15 mg kg-1 (Kabata-Pendias 2011), 18 mg kg-1 (Eriksson 2001), 19 mg kg-1 (Burt et al. 2003), 22 mg kg-1 (Licht 2005), 23 mg kg-1(John 2011), 25 mg kg-1 (FOREGS 2005) and 28 mg kg-1 (Takeda et al. 2004). According to Kabata-Pendias (2011), Pb in surface soils varies from 0.6 to 65 mg kg-1. On the other hand, Pais and Jones (1997) reported that natural levels of Pb in soils remained below 50 mg kg-1. Lead concentration in rocks varies greatly with their types. Kabata-Pendias (2011) reported that range and mean values of Pb in different rocks were respectively 2.3–70 and 22 mg kg-1 for podzols, 1.5–70 and 28 mg kg-1 for cambisols, 10–50 and 26 mg kg-1 for calcisols/leptisols, 8–70 and 23 mg kg-1 for kastanozems /chernozems, and 1.5–176 and 44 mg kg-1 for histosols. Therefore, the background Pb level in soils varies greatly with the type of parent material from which it is developed.
Although, the natural background levels of Pb in soil remain below 65 mg kg−1 (Kabata-Pendias 2011), some authors reported high levels of soil Pb as a result of human activities. Arshad et al. (2008) reported 1830 mg Pb /kg soil in Bazoches (North-Western France), and 39250 mg Pb /kg soil in urban areas of Toulouse (South-Western France). Shahid et al. (2013b) reported 2914 mg Pb /kg soil near Evin-Malmaison (North of France). Similarly, 4151 mg Pb /kg soil was reported by Austruy et al. (2014) in soil samples collected from 500 m North-East of former Metaleurop Nord smelter located in Northern France. In Baltimore, Pb level in garden soil was recorded as 10,900 mg kg-1 (Mielke et al. 1983) and in Chicago Pb contamination in residential soil has been found as high as 7,950 mg kg-1 (Shinn et al. 2000). Aliu et al. (2016) reported 35000 mg kg-1 of Pb in soil near Pb-Zn mining and smelting area of the Mitrovica Region, Kosovo. Lead does not undergo biological or chemical degradation and persists in both water and soil for a long time. It is reported that Pb can be reserved in environment for about 150 to 5000 years (Saxena et al. 1999). During Pb-acid battery recycling, different steps of the
3.0
3.5
4.0
4.5
5.0
5.5
10
3T
ho
us
an
ds
me
tric
to
ns
Lead
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process (fusion, crushing, refining, reduction etc.) emit various Pb species such as anglesite (PbSO4), metallic lead (Pb0), cerussite (PbCO3) and lead oxide (PbO) (Uzu et al. 2011).
Increase in Pb level in soil seems because of its various natural and anthropogenic sources. Natural sources of Pb include erosion and weathering of parent rocks containing high levels of Pb in their structure (Table 3.1). As a result, high levels of Pb are transferred to water and land. Volcanic eruptions also add significant amount of Pb in the environment. For example, about 540 to 6,000 tons of Pb was emitted from volcanoes in 1983 (Nriagu and Pacyna 1988) and about 110 x 106 tons in 2001 (Richardson et al. 2001). In 2015, about 2,530x106 metric tons of Pb has been mined from the Earth’s crust (US-GS 2016). About 4- to 10x105 tons of Pb was discharged with waste released from metal mining sites in 1983 (Nriagu and Pacyna 1988).
Table 3.1: Lead concentrations in different constituents of Earth's crust
Earth's crust constituents Pb level (mg kg−1) Reference
Basaltic igneous 2-18 Cannon et al. 1978
Granite igneous 6-30 Cannon et al. 1978
Shales and Clays 16-50 Cannon et al. 1978
Black shales 7-150 Cannon et al. 1978
Sand stone <1-31 Cannon et al. 1978
Soil 2-100 Allaway 1968
Agriculture Crops 0.1-10 Allaway 1968
Main two anthropogenic sources of Pb for urban soils are Pb-based paints and leaded gasoline. Lead was used in paints which resulted 4 to 5 million metric tons of Pb discharge into the atmosphere (ATSDR 2007b). According to estimation, 24 million of US housing units have a significant Pb based paint hazard (Jacobs et al. 2002). Lead has been used in gasoline since start of 1920s and was decreased in 1986 significantly. Use of Pb in gasoline is banned in the US as fuel additive since 1996 (Kerr and Newell 2003). On-going Industrialization, rapid urbanization and land use change are deteriorating the soil quality in urban areas rapidly with respect to Pb and other metals levels. Moreover, direct dumping of industrial and urban waste, atmospheric depositions, agricultural runoff and discharge of untreated effluent adversely affect urban soil quality with respect to Pb contents (Sharma and Dubey 2005).
3.4 PHYTOAVAILABILITY OF LEAD IN SOILS
The bioavailable fraction of a metal in soils is of utmost importance in order to predict the environmental risks related with Pb contaminated sites. Lead is one of those most important and major environmental contaminants having limited solubility in soils and its plant availability is minimal due to complexation with organic matter, clays, sorption onto oxides and precipitation as hydroxides, phosphates and carbonates (Shahid 2010). In fact, soil properties greatly influence metal behaviour in soil-plant system (Hussain et al. 2006, 2007). Owing to its low mobility and capacity to form complexes with soil constituents, most of Pb concentrates in top 25 cm soil (Cecchi et al. 2008). Now-a-days, it is well-known that total Pb contents in soil did not fully
56
represent its mobility, phytoavailabiltiy and toxicity. This is because of the fact that different chemical forms (chemical speciation) of Pb, and their varying adsorption/desorption in soil influence its behaviour in soil-plant system (Saifullah et al. 2015). Lead is reported to occur in various forms in soil such as (i) free metal ions, (ii) complexed with inorganic anions (CO3
2–, HCO3–, Cl– and SO4
2–) or complexed with organic ligands (amino acid, fulvic acid and humic acid). In addition to these forms, Pb may also occur as attached /adsorbed onto particle surface of Fe-oxides, biological material and organic matter and clay particles (Vega et al. 2010). Lead also exists in mineral forms that are highly stable with low solubility in soil, e.g. hydrocerussite (Pb3(CO3)2(OH)2), litharge (PbO) and cerussite (PbCO) (Essington et al. 2004). Pyromorphite (Pb5(PO4)3Cl) is another stable Pb form in soils with high contents of phosphate (Scheckel et al. 2005).
Owing to its nature of frequent and strong binding and complexation, very small amount of Pb in soil is mobile, soluble and phytoavailable (Pourrut et al. 2011). Behaviour of Pb in soils in terms of its mobility, solubility and phytoavailability is governed by a complex interaction of various factors (Fig. 3. 2) (Shahid et al. 2012a). These factors include soil mineralogy, soil pH, redox conditions, microbial and biological conditions, cation-exchange capacity, level and form of Pb present in soils, level of competing cations, levels of organic and inorganic ligands and plant species (Shahid 2010). These factors act individually and in combination with each other to modify soil Pb behaviour as well as phyto-absorption rate and level. Lead phytoavailability is greatly dependent on its chemical species and by the level of Pb as free ions (Shahid et al. 2011, 2014b).
3.5 SOIL PROPERTIES AND LEAD PHYTOAVAILABILITY
Mobility pattern for of metals in soils differs because of differences in soil parameters. In soils, availability of Pb depends on the path in which they partition between solid phase and soil solution. Heavy metals fractionation in soils, their mobility and accessibility, prospect of metals removal with washing methods are measured by soil properties such as organic matter, soil texture, type of clay minerals and content and Mn, Al, Fe oxides l, and physicochemical conditions (pH, aeration, saturation, redox potential).
3.5.1 Soil pH and Lead Phytoavailability
Soil pH refers to the acidic or alkaline nature. Previous studies have shown that soil pH is an important factor in controlling Pb adsorption onto soil constituents (Saifullah et al. 2010; Murtaza et al. 2011). Soil pH is important for Pb2+ free activity in soil solution as most precipitation reactions and adsorption of Pb are favoured by high pH (Shahid 2010). With increasing pH of soil solution, there is also an increase in negative surface affinity (CEC) of soil for metal ions such as Pb2+ (Wu et al. 2003). An increase in pH (basic range) results in higher adsorption of Pb to soil particles and a decrease in the absorption of Pb and other metals (Cd and Zn) by plants (Kuo et al. 1984). Soil pH also affects chemical speciation of Pb in soils, and thereby the biogeochemical behaviour in soil-plant system. Shahid et al. (2012d) demonstrated that free ionic Pb2+ dominates (80-90%) under acidic soil (pH < 6) conditions, whereas Pb-OH (90-100%) is the dominant Pb species under alkaline conditions (pH > 9). A similar pattern of Pb speciation was described by Yang et al. (2010) while reporting its adsorption on β-MnO2. Soil pH affect metal geochemical behaviour in soil-plant system through altering the solubility of carbonates, hydroxides and phosphates fractions, ion-pair formation, hydrolysis, surface charge of Fe and Al oxides as well as organic matter solubility (Romero-Freire et al. 2015).
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`
Figure 3.2: Biogeochemical behaviour of lead in soil-plant system
Lead is mostly associated with carbonate soil fractions and organic matter (Kabata-Pendias and Pendias 2011). Solubility and activity of Pb in slightly to moderately contaminated soils are controlled by Pb2+ ions, its strong adsorption with soil minerals (particularly oxides of Fe) and organic matter (Gustafsson et al. 2011; Romero-Freire et al. 2015). Surface complexation model depicts that adsorption of Pb is dominated by organic matter contents at a soil pH < 6 whereas Fe oxides adsorption is more prevalent at a pH > 6 (Gustafsson et al. 2011). The presence of organic matter in soil especially humic substances is highly important for mobility and speciation of Pb because of its high functionality (amino, carboxyl and hydroxyl groups) and thereby modifies the phytoavailability, adsorption, retention, and migration of Pb in natural environment. Decrease in Pb mobility in soils with time seems due to ageing processes and adsorption is mainly governed by organic fractions of soil (Shahid et al. 2012a). However, these processes are reversed by change in soil physicochemical properties and rhizosphere environment.
Pb sequestration
in roots
Pb translocation
to shoots
ROSO2
H2O2
Pb Detoxification
Glutathione, Vacuolar
sequestration, Phytochelatins,
Antioxidant enzymes (CAT,
POX, LOX, SOD, GPX, APX)
Free metal ions or complexed with
organic and inorganic forms
CO32–, HCO3–,Cl– & SO4
2–
Cyclic nucleotide
gated ion channels,
Cation transporters
OH
Adsorbed Pb Pb2+
Pb in soil
Pb uptake
Pb Toxicity
Affect photosynthetic apparatus
and pigments, reduced plant
growth, genotoxicity, disturbed
water status, impaired nutrient
uptake, lipid peroxidation, toxic
effect on DNA and RNA
Sources of Pb in soil
Dumping of industrial &
urban waste, agricultural
runoff, untreated effluent,
paint & leaded gasoline
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3.5.2 Soil Organic Matter and Lead Phytoavailability
Lead is immobilised by organic matter through formation of stable complexes like metal-humus. Dissolved organic matter can bind Pb2+ ions and thus could decrease its available concentration in soil (Quenea et al. 2009; Romero-Freire et al. 2015). Halim et al. (2003) showed that application of humic acid (2%) significantly decreased exchangeable and soluble Pb forms. Similarly, Farrell et al. (2010) reported a decrease in soil solution Pb contents after the humic substances application as compost. Botero et al. (2010) reported that humic substances have high binding strength for Pb compared to other essential nutrients such as Ca, Fe and K. On the other hand, dissolved organic matter forms organo-Pb complexes that increased solubility of Pb as pH increased above 6.5 (Shahid et al. 2012d). In fact, Pb in non-acidic soil occurs as Pb-organic matter complexes (Ashworth and Alloway 2008). Evangelou et al. (2004) reported that addition of humic substances enhanced Pb bioavailable portion by forming soluble Pb-humic substances complexes. Soil organic matter can also enhance Pb availability for plants by enhancing its solubility, soil cation exchange capacity and providing metal chelates in soil solution (Saifullah et al. 2009; Romero-Freire et al. 2015). This dual role of organic matter towards Pb solubility and mobility depends upon the organic matter composition (Shahid et al. 2012a). In different studies, authors discussed the organic fraction role in the phytoavailability of metals in organically amended calcareous soils (Ghafoor et al. 2008; Shahid et al. 2012a), but only limited information is available on soils with natural organic matter content gradient.
3.5.3 Soil Microbial Activity and Lead Phytoavailability
Microbes in soils are one of the key indicators for soil quality. Soil microbes are involved in important geochemical processes which govern the fate and behaviour of essential nutrients and metals in soils (Leveque et al. 2014; Ahmad et al. 2016). These processes include decomposition of organic matter and transformation of phosphorous (acid phosphatase), nitrogen (urease) and carbon (invertase). Soil microbial activity is also an indicator of heavy metal pollution because soil enzymatic and microbial activities are sensitive to the contamination of heavy metals including Pb. Lead can greatly influence soil microbial activities by decreasing their biomass, diversity and the structure of community (Renella et al. 2005). It has been described that Pb decreased activities of enzymes (CAT, urease, invertase, acid phosphatase) significantly in soils (Belyaeva et al. 2005).
Microbes are known to increase or decrease Pb phytoavailability by changing its adsorption/desorption in soils. Soil amendment with heavy metals-solubilizing microbes such as plant growth-promoting bacteria (PGPR) is a promising approach to increase phytoavailabiltiy of Pb and other metals in soils. Inoculation with rhizosphere microorganisms caused an increase in the phytoavailability and accumulation of Pb, Cd and Zn by Sedum alfredii (Xiong et al. 2008). Similarly, microbial inoculation of soils enhanced the phytouptake of Pb, Cd and Zn by Thlaspi caerulescens (Epelde et al. 2010). Arbuscular mycorrhizal fungi have also been reported to enhance absorpton of Pb, Cd, Zn and Cu by Litsea cubeba, Perilla frutescens, Phytolacca Americana, Dysphania ambrosioides and Rehmannia glutinosa (Long et al. 2010). In contrast to above mentioned results, some studies reported a decrease in Pb phytoavailability when inoculated with microorganisms. For example, Dary et al. (2010) reported that inoculation with Brady rhizobium significantly decreased absorption of Pb, Cd and Cu by Lupinus luteus.
In soils, microorganisms affect fate and behaviour of Pb and other pollutants directly or indirectly. Microbial activities in soil change physicochemical characteristics of the environment and therefore indirectly may affect the mobility and speciation of Pb. In soils, microbial activities can strongly affect physical and chemical soil characteristics (pore space, nutrients availability, redox potential, acidity etc.). Similarly, soil microbial activities also affect biogeochemical processes in soils, which in turn affect the fate and behaviour of trace elements i.e. their mobility, solubility, toxicity and phytoavailability (Sessitsch et al. 2013).
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3.6 SOIL-PLANT TRANSFER OF LEAD
3.6.1 Molecular understanding of Lead Bio-absorption
Although, Pb is immobile element in soil due to formation of stable and immobile complexes with soil constituents, but a soluble and exchangeable fraction of Pb is absorbed by plants and other living organisms. In this way, soils contaminated with Pb can contaminate the food chain. Lead enters plants from soils mainly through root absorption (Uzu et al. 2011). Soil-plant transfer of Pb is not direct but occurs in steps. Firstly, Pb present in soil solution is adsorbed onto plant roots (Pourrut et al. 2013). Several studies have reported Pb adsorption onto root cells of different plant species such as Funaria hygrometrica, Brassica juncea, Festuca rubra, Vigna unguiculata and Lactuca sativa (Pourrut et al. 2011 and references therein). The Pb then binds to polysaccharides of the rhizoderma cell surfaces or to carboxylic group of mucilage uronic acid (Seregin and Ivanov 2001). After adsorption onto root cells, Pb generally enters root cells passively and moves through water stream (Pourrut et al. 2011). The adsorption of Pb onto plant roots does not take place uniformly due to variation in Pb concentration (Seregin et al. 2004). Generally, Pb adsorption and entrance into roots is more in roots apices compared to other parts. Indeed, root cells are comparatively young in root apices with thin cell walls that do not block root absorption of Pb (Seregin et al. 2004). Furthermore, rhizodermic pH is the lowest in apical area, which enhances Pb solubility in soil solution.
Lead can enter into root cells through several pathways. Generally, H+/ATPase pumps facilitate Pb absorption by roots by maintaining a high negative membrane potential in rhizoderm cells (Wang et al. 2007). Calcium is well-known to block Pb absorption by plant roots (Pourrut et al. 2008). Calcium-induced decrease in Pb absorption is mostly due to competition between these two cations for Ca channels (Pourrut et al. 2008). Once Pb enters into roots, most of it binds with ion exchangeable sites in cell walls and the extracellular precipitation as carbonate and phosphate compounds (Sharma and Dubey 2005). The unbound form of Pb is transported through Ca channels and accumulated near endodermis (Pourrut et al. 2008). Inhibition of the Pb uptake by calcium is well-recognized (Pourrut et al. 2011). Numerous studies reported that Ca permeable channels in plants are main pathways through which Pb enters into root cells (Pourrut et al. 2008). In addition to Ca channels, Pb can enter into roots via non-selective pathways, such as low affinity cation transporters (Wojas et al. 2007) or cyclic nucleotide-gated ion channels (Kohler et al. 1999).
Plants such as Zea mays, Pisum sativum, Hordeum vulgareand, Brassica juncea were recently focused for the Pb phytoremediation research. Plant activities and characteristics have been reported to affect Pb availability in soils and its absorption by plants. These characteristics include mycorrhization, root surface area, root exudates, transpiration rate etc. Different species of plants showed diversity in root activities and root exudates which in turn affected the availability and solubility of Pb in soils (Shahid et al. 2012a). Absorption of Pb is affected greatly by rhizospheric processes. Lin et al. (2004) discussed the ability of Oryza sativa to acquire higher Pb levels from soil secreting organic acids (root exudates) and thereby decreasing soil pH. Phytovailability of Pb increases with a decrease in soil pH. Moreover, Pb complexes with organic acids secreted by plant roots have comparatively more phytoavailability compared to free ions. Lin et al. (2004) also reported that concentration of extractable Pb in the rhizosphere soil was higher than bulk soil demonstrating the role of root activities towards Pb solubilisation and phyto-absorption.
3.6.2 Lead Sequestration in Plant Roots
After absorption, Pb may accumulate in root cells or may be transferred to shoots. In majority of plant species, major portion (≥ 95% or more) of absorbed Pb is sequestered in root cells with limited translocation to shoots. According to autoradiographic and ultrastructural studies conducted on Pb localization and translocation in plant roots (Pourrut et al. 2011), deposition of Pb in ground meristem tissues is high and non-uniform. Lead sequestration in root cells has been
60
demonstrated in several plant species such as Funaria hygrometrica, Festuca rubra, Brassica juncea, Vigna unguiculata, Lactuca sativa, Allium sativum, Vicia faba, Vigna unguiculata, Sedum alfredii and Avicennia marina (Pourrut et al. 2011; Shahid 2010 and references therein).
Although sequestration in plant roots has been reported for several metals, but this phenomenon is very intense and specific for Pb. Different researchers proposed several reasons and mechanisms behind it. These reasons include: accumulation in vacuoles of root cells, precipitation of Pb as insoluble salts in roots, blockage by Casparian strip of endodermis, sequestration in plasma membranes and immobilization by negatively charged pectins (Sharma and Dubey 2005; Shahid 2010). Within the nucleus, Pb binds with orthophosphate ions (Tandler and Solari 1969). Lead is found in cell cytoplasm in association with electron dense precipitates present in vesicles or organelles and membranous inclusions. Lead is not translocated apparently into reproductive/fruit tissues. Lead moves predominantly into the apoplast of roots and accumulates near endodermis. For Pb movement from roots to shoots, endodermis behaves as partial barrier (Shahid et al. 2014b). Distribution, transportation and accumulation of Pb in different plant parts vary at different plant growth stages (Pourrut et al. 2011).
Sub-cellular accumulation of Pb in plants depends mainly on plant species as well as genotypes and even cultivars of same species. Generally, tolerance and detoxification mechanisms inside plants mediate Pb sub-cellular accumulation. Some plant species can endure Pb toxicity through inactivation and complexation (Hordeum vulgare, Allium cepa, Zea mays) and other show toxicity (Phaseolus vulgarisand Brassica napus) due to well-known toxicity of Pb to metabolic processes. According to some researchers, enhanced accumulation of Pb in root cells is a part of tolerance mechanism. Phytochelatins are produced in roots exposed to Pb suggesting their role for detoxification of PCs (Pourrut et al. 2011).
3.6.3 Lead Translocation from Roots to Shoots
Lead transfer from roots to shoots takes place through xylem loading driven by transpiration (Verbruggen et al. 2009), with apoplastic and/or symplastic transport. However, this translocation of Pb from roots to shoots is impeded apparently by several physical and biochemical processes involved in Pb binding, precipitation and/or inactivation. Using X-ray mapping, Arias et al. (2010) showed high deposition of Pb in phloem and xylem cells of Prosopis glandulosa. It is reported that after absorption, Pb penetrates into the central cylinder of stem, and can again take apoplastic pathway for translocation. Vascular flow translocates Pb towards leaves (Krzesłowska et al. 2010). However, inside xylem, Pb may bind with organic or amino acids (Maestri et al. 2010). Researchers generally use translocation factor (ratio of Pb in shoots to that in roots) to indicate the proportion of Pb translocation. Generally, value of translocation factor for Pb is low for most plant species due to its high sequestration into roots (Uzu et al. 2009).
It was reported that roots absorbed a large amount of Pb from soils while at the same time restricted its translocation to aerial parts. Lead was found in higher amount in cell walls, vacuoles, intercellular spaces and dictyosomes as a Pb pyrophosphates, crystals and precipitates (Qureshi et al. 1986). Some plant species can transfer and accumulate higher Pb in their shoots compared to roots (Arshad et al. 2008). Such hyperaccumulator plant species include pelargonium and Brassica pekinensis can transfer higher Pb concentrations to shoot tissues, without incurring any toxic effect to their biochemical and metabolic processes (Arshad et al. 2008). Generally, it is believed that a specific Pb hyperaccumulator can uptake > 1000 mg kg-1 Pb in their plant parts (Maestri et al. 2010). These plants are reported to uptake and translocate high levels of Pb by some specific characteristics; secretion of root exudates which can solubilize and enhance Pb uptake by roots (Arshad et al. 2008) as well as chelation of Pb by ligands and their transportation to shoots. Moreover, these hyperaccumulator plant species tolerate and detoxify high concentrations of Pb because they have well-developed homeostasis mechanisms such as selective and/or limited metal uptake, excretion from root cells, binding with specific ligands, and sequestration in cell organelles.
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3.7 TOXIC EFFECTS OF LEAD
Lead is non-essential for plants and its biological functions in plant metabolism are still unknown. Excess of Pb causes negative effects in plants. At cellular levels, Pb affects hormonal status of plants and it impedes membrane permeability (Fig. 3.3, Ann. Table 2). At higher applied levels, Pb negatively affect seed germination, plant growth and development, and causes chlorosis in the leaves and blackening of the root system (Islam et al. 2008). Lead toxicity in plants can cause change in the catalytic activities of different enzymes, disturbance of mineral nutrition, decreases photosynthetic pigment contents and increases in membrane permeability (Sharma and Dubey 2005). These toxic effects result in inhibition of plant growth and biomass production. In addition, Pb stress can also alter genes expression. Lead also upsets mineral nutrition, water balance and inhibits photosynthesis (Pourrut et al. 2011).
Its higher concentration affects photosystem II functioning and damages its structure and thus could decrease absorbance of visible light (Cenksi et al. 2010). Lead can inhibit the activity of enzymes and causes disruption of plants protein by binding with sulfhydryl groups of several proteins. It could cause displacement of elements in plant cells causing deficiency effects (Pourrut et al. 2011).
3.7.1 Lead Toxicity to Plant Growth
In plants, excess amount of Pb causes toxicity symptoms associated with retarded root elongation, growth, enzyme activities and appearance of leaf chlorosis (Sharma and Dubey 2005). Generally, Pb could decrease seed germination, even at very low concentration, in Zea mays, Elsholtzia argyi, Hordeum vulgare, Pinus halepensis, Spartiana alterniflora and Oryza sativa (Islam et al. 2007), and Zea mays (Islam et al. 2007; Sengar et al. 2009). However, high levels of Pb may speed up seed germination process, and simultaneously causes deleterious effects on the length of hypocotyl and radical in plants (Islam et al. 2007). Lead-induced decreased seed germination can be due to interaction of Pb with amylase and protease enzymes (Sengar et al. 2009).
Figure 3.3: Toxic effects of lead accumulation in plants.
Lead causes reduction in roots growth and results in mitotic irregularities in plants (Shahid et al. 2011). Germination of seed is an initial event for plant life and initiates with regulation of enzymatic reactions that causes activation of anabolic and catabolic processes in embryonic axis and storage tissues, respectively. Breakdown of storage tissues is one of the
Toxic effects of Pb
in plantsPhotosynthetic
apparatus and pigment
contents
Nutrient & water
uptake
Decreased plant growth &
yield
Biochemical & metabolic
changes
Genotoxicity
(RNA & DNA damage)
Enhanced ROS
production & oxidative
damage
62
crucial events resulting seed germination following water imbibition. According to Ekmekci et al. (2009), Pb ions cause water deficiency by disturbing water balance leading to growth reduction in plants. It is supposed that decrease in plant growth might be related to the mitotic index inhibition with treatment of Pb metal (Shahid et al. 2011). Similarly, at increasing Pb concentration, fresh and dry biomass of roots, shoots and leaves was affected negatively in plants (Pourrut et al. 2011). Atici et al. (2005) reported that metals, especially Pb, caused inhibitory effects on seedling growth and seed germination.
For most plants species, Pb concentration > 30 mg kg-1 in plant tissue is considered toxic (Ruley 2004). Main effect of Pb toxicity is quick inhibition of the root cell growth perhaps due to blockage of cell division in root tips. In several plants species, including Zea mays, Triticum aestivum, Sedum alfredii and Pisum sativum, Pb-induced decrease in root length and plant biomass has been demonstrated. Verma and Dubey (2003) recorded decrease in root length in Oryza sativa seedlings up to 40% by Pb in 20 days old seedlings. Furthermore, after exposure to different Pb concentrations, cell wall distention, formation of folds, nicks and protuberances were found in Elsholtzia argyi, Allium cepa and Triticum aestivum (Pourrut et al. 2011). Jiang and Liu (2010) observed vacuolization of dictyosomes and endoplasmic reticulum, loss of cristae, deep coloured nuclei, mitochondrial swelling and injured plasma membrane to the root of Allium sativum after 48–72 hr of Pb exposure.
3.7.2 Lead Genotoxicity
Lead is known to induce antimitotic effects in plants. Long ago Hammett (1928) showed that Pb inhibited mitotic activity in root cells of Allium cepa. Later, Patra et al. (2004) demonstrated that Pb can prolong interphase and shorten the mitotic stage, thereby increasing the duration of cell cycle. Lead binds to cell walls and membranes, which causes rigidity in these tissues and decreases cell division. After that, Pb disrupts microtubules that play a key role in mitosis. Lead generally disturbs the M and G2 phases of cell division that causes generation of abnormal cells during colchicine-mitosis stage. Lead-induced toxicity to cell cycle can be due to direct or indirect Pb interaction with proteins (such as cyclins) involved in cell cycle. Lead does not cause any effect on mitosis and cell division at low level of exposure. However, Pb-induced chromosomal aberrations, micronucleus formation and chromosomal fragmentation can occur even at low Pb exposure of plants (Cecchi et al. 2008). Lead can induce chromosomal aberrations by disrupting the microtubule network. Lead affects DNA-protein or horizontal DNA-DNA links by causing disruptions in single and double strands of DNA (Gichner et al. 2008). Similarly, Shahid et al. (2011) reported that Pb can disturb mitotic index and cell division in Vicia faba root tips.
Application of heavy metals, especially Pb at DNA level, can induce genotoxicity (Shahid et al. 2011) and its ions form covalent bonds with the DNA sequence (Chaoui et al. 1997). Various DNA assays have been used to determine the genomic DNA alterations induced by Pb and other heavy metals in plants (Ahmad et al. 2012). Cenksi et al. (2010) showed that Pb amplifies DNA fragments of the genomic DNA. They also reported that Pb affected stability of genomic template in Brassica rapa. After binding with DNA, Pb destroys DNA replication and repair mechanisms. Lead does not cause direct genotoxic effects unless it is bound to the naked DNA (Valverde et al. 2001). Lead strongly binds to the several enzymes, RNA, DNA and amino acids, and thus it disturbs several metabolic pathways. Lead can replace Zn that also affects DNA replication in Zn finger pattern of enzymes and impairs the DNA repair (Gastaldo et al. 2007). Lead is reported to cause genotoxic effects in various species of plants including Sesbania drummondii (Srivastava et al. 2007), Oryza sativa (Verma and Dubey 2003), Brassica juncea and Sesuvium portulacastrum (Zaier et al. 2010), Triticum aestivum (Ekmekci et al. 2009), Chenpodium album and Salsola passerine (Hu et al. 2012). Lead-induced genotoxic damage varies with plant species. Hypeaccumulator plants (Sesbania drummondii, Sesbania grandiflora) have potential for various heavy metal hyperaccumulation. Lead exposure of Sesbania grandiflora plant did not cause any genotoxicity. This can be due to plants capability for activation of their defensive mechanism against genotoxicity and oxidative damage by Pb ions. Samardakiewicz and Wozny (2005) showed that cell division in roots of Lemna minor was more
63
sensitive to Pb compared to those of terrestrial species such as Tradescantia, Zea mays and Vicia faba.
3.7.3 Lead-Induced Oxidative Stress
The generation of ROS under aerobic conditions is an unavoidable feature of life (Foyer and Noctor 2012). Different plant organelles (peroxisomes, chloroplasts and mitochondria) are major places of ROS generation in plant cells (Shahid et al. 2015b). Electron transport chain of mitochondria (complex I and III) is the main site of ROS production in plants (Sabir et al. 2015). Almost 2% of the total O2 utilized by mitochondria is converted to H2O2 by different biochemical reactions in plants. Like other heavy metals, Pb exposure can induce over-production of ROS in different plant cells. Lead exposed plants suffer breakdown of various metabolic processes as a result of oxidative stress. Lead can produce different ROS such as hydrogen peroxide (H2O2), superoxide radicals (dO2) and singlet oxygen (1/2O2) in plant cells (Baxter et al. 2014). Lead-induced over-production of ROS can cause membrane damage and lipid peroxidation, and elicits oxidative stress (Khan et al. 2016). Over-production of ROS can rapidly attack proteins, nucleic acids, amino acids and lipids.
Lead may induce ROS generation interacting with oxyhemoglobin which undergoes autoxidation. Lead-induced over-production of ROS can cause significant modification in the lipid composition of cell membranes (Shahid et al. 2015b). The ROS have potential to severely affect the esters and poly-unsaturated fatty acids of lipids (Mishra et al. 2006; Pourrut et al. 2013). Lead-induced lipid peroxidation in cells occurs by the removal of hydrogen from unsaturated fatty acids via ROS interaction that forms reactive aldehydes and lipid radicals (Mishra et al. 2006). Lead-induced increase in ROS has been reported in various plant species such as Sesbania grandiflora, Arabidopsis thaliana, Ceratophyllum demersum, Zygophyllum fabago and Vicia faba (Malar et al. 2014; Qi et al. 2015; Chen et al. 2015; Lopez-Orenes et al. 2014; Khan et al. 2016).
64
Table 3.2: Lead-induced enhanced production of ROS in various plant species grown in
hydroponic conditions
ROS Plant species Pb exposure level Time duration References
H2O2, O2- Vallisneria natans 25, 50, 75,100 µM 4 days Wang et al. 2011
H2O2 Spinacea oleracea 2.42, 4.83 mM 10 days Khan et al. 2016
H2O2, O2- HypnumPlumaeforme 0.01, 0.1, 1, 10 mM 2 days Sun et al. 2011
H2O2, O2- Sesbania grandiflora
100, 200, 400, 600, 800,1000 mg L-1
10 days Malar et al. 2014
H2O2 Arabidopsis thaliana 500 µM 7 days Qi et al. 2015
H2O2, O2-
Ceratophyllum demersum
5, 10, 20, 40, 80 Mm 7, 14, 21 days Chen et al. 2015
H2O2 Zygophyllum fabago 0.75 mM 7 days Lopez-Orenes et al. 2014
H2O2, O2- Vicia faba 5 µM 1, 4, 8, 12, 24 h Shahid et al. 2014d
H2O2 Talinum triangulare 0.25, 0.5, 0.75, 1.0,1.25 mM
1, 3, 5, 7 days Kumar et al. 2012
H2O2 Triticum aestivum 1, 2, 4 mM 6 days Yang et al. 2010
H2O2 Nasturtium officinale 50, 100, 200, 250, 500 mg/l
2 days Keser and Saygideger 2010
H2O2, O2- Lathyrus sativus 0.5 mM 4 days Brunet et al. 2009
H2O2, O2- Sedum alfredii 200 µM 1 day Huang et al. 2008
H2O2 Peganum harmala 5, 10, 25, 50 mg/l 14 days Mahdavian et al. 2016
H2O2 Prosopis farcta 80, 160, 320, 400, 480 μM 12, 24, 48, 72,
96 h Zafari et al. 2016
H2O2, O2- Nicotiana tabacum 100, 500µM 3 days Alkhatib et al. 2013
H2O2 Brassica napus 50, 100µM 42 days Shakoor et al. 2014
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide.
65
3.8 LEAD DETOXIFICATION
In order to survive and cope with metal-induced stress, plants are equipped with an effective natural defence system (Figure 3.4). Plants defence system helps them to survive under metal stress by inhibiting tissue dysfunction and cell injury (Shahid et al. 2014e). Plants defence system contains a number of molecules, ligands and enzymes that operate separately or in combination with each other to detoxify over-production of ROS. However, the intensity and activation of a specific defence system vary with plant maturity and type, as well as the length of metal exposure. It is general understanding that compared to plants that are sensitive to heavy metals, stress-tolerant plant species have a more effective defence system to protect them against ROS injury (Liu and Pang 2010). As a result, these plants can accumulate and tolerate high levels of Pb and other toxic metals without experiencing any major phytotoxic effect (Rascio and Navari-Izzo 2011).
Plants have different defence mechanisms to deal with the heavy metal toxicity. These defence mechanisms operate at different stages of metal entrance to plants. The primary and first defence strategy adopted by tolerant plants is the decreased uptake and accumulation of Pb in roots. Plants secrete several types of exudates which help plants to avoid metal accumulation. Lead adsorption on plant roots is considered a defence mechanism adopted by plants to decrease its uptake by plants, e.g., binding of Pb with pectin carboxyl groups in root cells is considered a defensive approach of plants to resist Pb entrance and toxicity (Jiang and Liu 2010). Secondary defence mechanism is comprised of several antioxidants to scavenge Pb-induced over-production of ROS (Khan et al. 2016). These antioxidant substances include GSH, APX, CAT, vitamin E and C, SOD and carotenoids (Pourrut et al. 2011). These substances are found in the cell components (organelles) especially in chloroplasts and vacuole. Scandalios (2005) suggested the following dissemination of antioxidant substances in plants: about 73% in vacuoles (peroxidase, glutathione and ascorbate), 17% in the chloroplasts (dehydroascorbate reductase, ascorbate, α-tocopherol, carotenoids, GR, APX, MDHR, glutathione and Cu/Zn-SOD); 5% in cytosol (CuZn-SOD, APX, CAT, POD, MDHR, GSH an GR); 4% in apoplast (POX and ASC); 1% in mitochondria (monodehydroascorbate radical reductase, Mn-SOD, GR, GSH and CAT) and peroxisomes (Cu/Zn-SOD, CAT).
3.8.1 Vacuolar Compartmentalization of Lead
Cellular compartmentalization of toxic metals is an important feature in metal tolerance and homeostasis. Some plant species can bind Pb to certain organic molecules, which is then transported to several plant cell compartments, like plasma tubules, endoplasmic reticulum vesicles, dictyosome vesicles or vacuoles (Pourrut et al. 2011).
3.8.2 Lead Chelation by Phytochelatins
Cellular compartmentalization of Pb is an important aspect for plant metal homeostasis. Phytochelatins (PCs) are regarded as one of the best documented mechanisms for cellular compartmentalization of toxic metals including Pb as well as translocation of essential nutrients. Phytochelatins are small, Pb and other metals-binding polypeptides having a general structural formula (γ-Glu-Cys) nGly (n=2-11). Phytochelatins bind Pb and other metals by mercaptide bonds through their functional groups and form Pb-PC complexes, thereby playing important role in their detoxification and homeostasis in plants (Verbruggen et al. 2009; Andra et al. 2009). It is proposed that PCs sequester Pb into the cytoplasm before transportation to chloroplasts and vacuoles (Shahid et al. 2014a). These complexes are then transferred to cell compartments such as chloroplasts and vacuoles. In this way, PC stores Pb in non-sensitive cell compartments and thereby decreases the toxic effect of Pb in cells (Andra et al. 2009; Scheidegger et al. 2011). Hyperaccumulator plant species have potential to bind and sequester metals in cellular compartments to inhibit their toxic effects. In these plants, metal homeostasis is achieved by binding Pb with ligands.
66
Figure 3.4: Lead-induced oxidative stress in plant cells
Exposure of plants to Pb and other heavy metals is demonstrated to activate PC synthesis and the phytochelatin synthase (PC-synthase) (Pourrut et al. 2011). Lead-induced increased PC production has been demonstrated in several plants including Phaeodactylum tricornutum (Scarano and Morelli 2002), Ceratophyllum demersum (Mishra et al. 2006) and Euglena gracilis (Mendoza-Cózatl et al. 2006). Phytochelatin synthase non-translationally catalyses synthesis of PC. This catalysis is triggered by metal ions such as Pb, Cd, Cu and Zn. During PC-synthase-mediated production of PC, previously synthesized PC and/or tripeptide GSH act as a substrate (Wojas et al. 2010). Lead-induced synthesis of PCs and activation of PC-synthase is documented in literature (Gupta et al. 2010), however, transportation and sequestration of Pb-PC complex to vacuole is not yet well-developed. It is reported that Pb and other metals do not activate PC-synthase directly, but indirectly by forming Pb-GSH complex. Lead blocks the thiol group by forming Pb-GSH complex which results in enhanced PC-synthase activity (Na and Salt 2010). Over-expression of PCs may cause depletion of GSH and thereby causes oxidative damage (Verbruggen et al. 2009).
Synthesis of PCs and its homologues is a distinctive defence mechanism in plants, algae and fungi. Over-expression of PCs after Pb exposure and formation of PC–Pb complexes are reported in literature (Verbruggen et al. 2009; Wójcik and Tukiendorf 2014). Gisbert et al. (2003) reported a significant increase in accumulation and tolerance of Pb following enhanced induction of Triticum aestivum gene (TaPCS1) in Nicotiana glauca that is considered responsible for encoding PC-synthase. After Pb exposure, PC production increased in roots and leaves of micro-algae (microphytes) (Mendoza-Cózatl et al. 2006) and Salvinia minima (Piechalak et al. 2002). In Salviniaminima, Pb accumulation induced changes in SmPCs gene expression which is responsible for PC synthesis. Several studies have reported changes in PC-synthase genes expression after Cd exposure but very rare data are available for Pb (Ramos et al. 2007). However, activation of PC and PC-synthase varies with plant organ. Vatamaniuk et al. (2000) reported rapid and intense activation of PC-synthase and production of PC in roots compared to leaves.
Pb2+
Pb-PC
Pb-GSH
Plasma
membrane
Vacuolar
compartmentation
H2O2O2•−
OH•
H2O
H2O
H2O
SOD APX
Halliwell–Asada
• Peroxisomes
• Chloroplasts
• Mitochondria
• Genotoxicity
• Lipid peroxidation
• Cell death
• Enzyme inactivation
• µnuclei induction
• Chromosomal aberration
NADPH
Lead
• HMAs
• ZIP
• ABC
• CDF
• Nramp
67
3.8.3 Lead Chelation by Glutathione
Glutathione, a sulphur containing tri-peptide, is a primary antioxidant group with low molecular weight. Glutathione is an important substance which governs the cellular redox status in cells (Pourrut et al. 2011; Rojas-Loria et al. 2014). Glutathione generally occurs in all the cell compartments such as mitochondria, endoplasmic reticulum, cytosol, chloroplast and vacuole. Glutathione has a general formula γ-glutamate cysteine-glycine. Synthesis of GSH is catalysed by two ATP-dependent enzymes; γ-glutamylcysteine synthetase (GSH1) and glutathione synthetase (GSH2). In Brassica juncea and Arabidopsis thaliana, GSH1 mainly occurs in plastids while GSH2 is confined to both cytosol and plastids (Wachter et al. 2005). One of the most important functions of GSH is to act as a substrate for PC synthesis, which is involved in metal detoxification (Rojas-Loria et al. 2014). In fact, Pb-induced activation of PC is via formation of Pb-GSH bond, which enhances GSH level in cells. Once GSH level is maintained in cells as a result of Pb stress, PC-synthase is activated thereby forming PC–metal complex. In this way, GSH plays a vital role in protecting plants against Pb-induced oxidative stress. Glutathione-induced protection of plants from Pb stress has been reported in several studies (Verbruggen et al. 2009).
In plants, GSH plays several key roles and participates in a number of cellular processes such as xenobiotics detoxification, sequestration of Pb and other toxic heavy metals (Freeman et al. 2004) and protection of plant cells against ROS stress (Foyer and Noctor 2005). In addition, GSH is involved in plant growth and developmental processes such as cell division and regulation of flowering (Ogawa et al. 2004). The GSH also acts as a co-factor and co-enzyme in the enzymatic detoxification reactions such as glutathione S-transferase (GST), gamma glutamyl-transpeptidase (GGT) and γ- glutathione peroxidase (GPX) (Foyer and Noctor 2005). Glutathione occurs in two different forms of oxidation; oxidized glutathione (GSSG) and reduced glutathione (GSH). The reduction potential of GSH depends on intracellular GSH/GSSG ratio (Foyer and Noctor 2005).
3.8.4 Antioxidant Enzymes and LeadToxicity
Lead is known to alter the activities of different antioxidant enzymes in plants in a specific manner (Khan et al. 2016). Antioxidant enzymes which are activated after Pb exposure include SOD, CAT, APX and POD (Table 3.3) (Pourrut et al. 2011). Different plants species show different tolerance and detoxification against Pb due to varying behaviour of antioxidant enzymes under Pb stress. Antioxidant enzymes scavenge ROS by donating an electron and neutralize ROS resulting in the formation of H2O. Previous studies reported Pb-mediated activation of enzymes in Brassica napus, Nasturtium officinale, Lathyrus sativus, Potamogeton crispus and Sedum alfredii (Keser and Saygideger 2010; Shakoor et al. 2014; Qiao et al. 2015).
Among the antioxidant enzymes, SOD is a key enzyme which is considered the first one involved in ROS scavenging. The SOD is a major O2
•− radical scavenger. Its enzymatic action results in O2 and H2O2 generation. The SOD removes O2
•− by catalysing the reduction of one O2•−
to H2O2 and oxidation of other one O2•− to O2 (Hasan et al. 2011). Lead-induced SOD activation
has been reported in Sesbania grandiflora (Malar et al. 2014), Suaeda heteroptera (He et al. 2015), Zygophyllum fabago (Lopez-Orenes et al. 2014) Sedum alfredii (Gupta et al. 2010). Lipo-xygenase (LOX) is ubiquitous in plants, and generates oxy-radicals and hydroperoxides by catalyzing peroxidation of the unsaturated fatty acid of bio-membrane (Aravind and Prasad 2003). The LOX and SOD activity represent extent of oxidative products. Activities of LOX and SOD were induced strongly in the presence of Pb in Sedum alfredii (Huang et al. 2008).
Catalase primarily occurs in mitochondria and peroxisomes and it decomposes H2O2 to H2O and O2. The H2O2 is toxic and is eliminated by converting it into H2O by CAT. Increase in activity of CAT can be explained with increasing its substrate (H2O2 level) as an adaptive system of plants (Shahid et al. 2014a). Activation of CAT under Pb stress is reported in different plant species, i.e. Ceratophyllum demersum (Chen et al. 2015), Oryza sativa (Jing et al. 2007) and Zea mays (Gupta et al. 2009). Guaiacol peroxidase, which occurs in cell walls, vacuoles, cytosol and extracellular spaces, is known as a stress-marker due to its high binding affinity for H2O2 and
68
great specificity for the phenolic substrates compared to CAT. The GPX consumes H2O2 and generates phenoxy compounds (Reddy et al. 2005). Increased activity of GPX enzyme under Pb stress represents high H2O2 generation (Verma and Dubey2003).
Peroxidases scavenge H2O2 by combining it indirectly with antioxidant compounds as guaiacol and ascorbate (Shahid et al. 2014a). The PODs are those antioxidant enzymes which catalyse H2O2 mediated oxidation of various substrates, especially phenolics (Saidi et al. 2013). Ascorbate Peroxidases are mainly present in cytoplasm and chloroplasts (Shigeoka et al. 2002). The APX is an important member of ascorbate-glutathione cycle (Gill et al. 2012), which mediates cellular redox status. The APX utilizes ascorbate for reduction of H2O2 to H2O and O2. An increase in APX activity is associated with H2O2 scavenging to protect plants against Pb damage. Three H2O2 scavenging enzymes (SOD, APX and GPX) generally exhibit simultaneous activation and decline due to their co-regulation (Shigeoka et al. 2002). Increase in APX activity upon exposure of Pb was found in Sedum alfredii (Gupta et al. 2013), Oryza sativa (Verma and Dubey 2003) and Vicia faba (Shahid et al. 2014e).
Glutathione Reductase functions in a chain of reactions known as ascorbate GSH cycle (Mittler 2002) to detoxify H2O2. The GR helps plant cells to maintain a high GSSG/GSH ratio (Noctor and Foyer 1998) that is vital to protect against oxidative stress. Increase in GR activity by Pb is also reported in Ceratophyllum demersum (Mishra et al. 2006), Vicia faba (Shahid et al. 2014e, 2014f) Vicia faba and Medicago sativa (Hattab et al. 2016). Dehydro-ascorbate reductase is an important antioxidative enzyme present in plants which mediates reduction of oxidized ascorbate to sustain a suitable ascorbate level in plants using GSH enzyme as reducing substrate (Nehnevajova et al. 2012). Ascorbic Acid (AsA) is the most abundant antioxidant enzyme found in plants and contributes in scavenging H2O2 through AsA-GSH cycle, which is constituted by APX, AsA, GR and GSH (Noctor and Foyer 1998). The AsA is mainly present in chloroplasts and is vital for plant defence against the oxidative stress (Noctor and Foyer 1998).
Lead-induced activation of antioxidant enzymes in different plant species grown in hydroponics
Enzymes Plant Species Pb exposure level Duration Reference
SOD, CAT, GPOX, APX Spinacea oleracea 2.42, 4.83 mM 10 days Khan et al. 2016
SOD, APX, GPOX,GR Vicia faba 5µM 1, 4, 8, 12, 24
hr Shahid et al. 2014f
POX, SOD, APX, CAT Sesbania grandiflora 100, 200, 400, 600,
800,1000 mgL-1 10 days Malar et al. 2014
POD, SOD, CAT Suaeda heteropteraφ 100, 200, 400, 600 mg kg−1 50 days He et al. 2016
POD, SOD, APX, CAT, GR, MDHAR, DHAR
Zygophyllum fabago 0.75 mM 7 days Lopez-Orenes et al. 2014
POD, SOD, APX, CAT Brassica napus 50, 100 µM 42 days Shakoor et al. 2014
SOD, AP, CAT, GR Nasturtium officinale 50, 100, 200, 250, 500 mg/l 2 days Keser and Saygideger 2010
GR, GST, APX Lathyrus sativus 0.5 mM 4 days Brunet et al. 2009
GK, OAT, ADC, DAO, PDH
Potamogeton crispus 25,50, 100, 200µM 5 days Qiao et al. 2015
SOD, APX, GHS, GR Sedum alfredii 5, 25, 100, 200 µm 1, 3, 5 days Gupta et al. 2010
POD, APX, CAT, NADH Vallisneria natans 25, 50, 75,100 µM 4 days Wang et al. 2011
69
POX, SOD, CAT, GR Brachytheciumb piligerum
0.01, 0.1, 1, 10 mM 2 days Sun et al. 2011
SOD, APX, CAT, GR, GPX, GSH
Ceratophyllum demersum
1, 10, 25, 50, 100 µM 1, 2, 4, 7
days Mishra et al. 2006
SOD, APX, GR, GSH Medicago sativa 10,100 µM 2, 7 days Hattab et al. 2016
POD, SOD, CAT Ceratophyllum demersum
5, 10, 20, 40, 80 µM 7, 14,21 days Chen et al. 2015
SOD,CAT, AsA Zea mays 25, 50, 100, 200 µM 1,4, 7 days Gupta et al. 2009
SOD, CAT, APX Oryzasativa 0.05, 0.15, 0.25 mmol/L 18 days Jing et al. 2007
LOX, SOD Sedum alfredii 200 µM 1 day Huang et al. 2008
SOD, APX, GR Medicago sativa 10, 100 µM 2, 7 days Hattab et al. 2016
GPX, APX, CAT, Suaeda fruticose 200, 400, 600 µm 90 days Bankaji et al. 2016
SOD, GR Oryza sativa 500, 100 µM 5-20 days Verma and Dubey 2003
POD, SOD, MDA Oryza sativaφ 89 mg kg−1 150 days Sun et al. 2015
POX, CAT Triticum durum 0.15, 0.25, 0.3 g/L 2 days Nedjah et al. 2013
SOD; superoxide dismutase, APX; ascorbate peroxidase, GR; glutathione reductase, POD; Peroxidase, LOX; lipoxygenase, GPX; guaiacol peroxidase, CAT; catalase, AsA; ascorbic acid,
GST; Glutathione S-transferase, AAO; ascorbic acid. φ Grown in soil culture.
3.8.5 Salicylic Acid and Lead Toxicity
Salicylic acid (SA), a phenolic hormone, is a known signalling molecule, which occurs naturally in plants (Khan et al. 2015). The SA is known to induce numerous effects in plants in biochemical and physiological activities. It plays an important role for the regulation of plants growth and its productivity (Hayat et al. 2010), regulation of seed germination, lowering glycolysis and fruit yield (Guo et al. 2009), stabilization of membrane and induction of gene expression (Belkadhi et al. 2015), regulation of nutrient contents under various stresses (Khan et al. 2015), maintenance of cellular redox homeostasis (Mateo et al. 2006), signalling defence response against several pathogenic infections, mediation of plant responses to abiotic stresses including heavy metal stress, temperature, UV radiation and salinity (Hayat et al. 2010; Khan et al. 2015). It is reported that SA decreases Pb-induced oxidative stress in plant cells and thereby retard its toxicity to biochemical and metabolic reactions taking place in plants (Tamás et al. 2015). For example, SA alleviated Pb-induced disruption of membrane in Oryza sativa seedlings has been reported (Chen et al. 2007).
Some studies showed that even exogenous application of SA induces beneficial effects and is involved to decrease Pb-induced generation of O2
−, H2O2 and lipid peroxidation in plants (Panda and Patra 2007). Overall, SA presence is considered to alleviate metal-induced toxicity in plants. The protective role of SA against Pb-induced toxicity has been demonstrated in Lemna mino (Kaur et al. 2013) and Oryza sativa (Chen et al. 2007). Although, SA is known to alleviate Pb-induced oxidative stress, however, the synthesis of SA and mode of action in plants are still not clear. The SA helps inhibit Pb absorption by plants. Lead absorption by plants occurs via cation channels such as those of Ca permeable channels (Sharma and Dubey 2005). During its movement through cation channels, Pb can bind with active binding sites of regulatory proteins
70
(Arazi et al. 2000). This showed that Pb absorption by plants is regulated with the regulation of cation channels. It suggests that Pb absorption and its binding with different proteins may play an important role in the molecular homeostasis of Pb (Kirberger and Yang 2008).
3.9 HORMETIC EFFECT OF LEAD TOXICITY
Hormesis is a bi-phasic process where low doses of a pollutant cause beneficial or stimulation effects but its high doses are toxic having detrimental effects on the biochemical and metabolic processes (Calabrese et al. 2007). Hormesis generally shows two types of curves, i.e., straight and inverted U-shaped curves, depending on the process evaluated (Fig. 3.5) (Jia et al. 2013). The former curve represents a decrease in the value of a parameter below control at the low level of application followed by an increase at the higher applied levels. On the other hand, an inverted U-shaped curve indicates an increase in the value of a parameter above control at lower doses following a decrease at the higher applied levels (Calabrese and Baldwin 2003).
For hormesis, straight U-shaped model generally denotes Pb toxicity parameters such as genotoxicity and mutagenesis. Under these circumstances, specifically low doses decrease toxicity by stimulating improvement compared to the control while higher applied levels above the NOAEL (no observed adverse effect level) could cause dysfunctioning. Inverted U-shaped curve shows normal functions such as longevity and growth, which is known as low dose induced stimulation and high dose induced inhibition models (Rodricks 2003). Wang et al. (2010) demonstrated a U-shaped dose-response curve under low doses Pb stress for activation of endo-proteinase isoenzymes, GPX and APX, production of O2
- radicals and generation of malondialdehyde in roots, indicative of decreased toxicity and oxidative stress. They also reported inverted U-shaped curves for SOD, growth height, inducible heat shock protein 70 and endo-proteinase activities with increasing Pb levels indicating enhanced toxicity and oxidative stress. Similarly, Yang et al. (2010) reported that low Pb level (1 mM) slightly stimulated seed germination and seedling growth of Triticum aestivum but at higher Pb levels, significantly decreased seed germination and growth parameters. Janota et al. (2015) reported that Pb-induced hormetic effect with respect to concentrations of indole-3-acetic acid and H2O2 at 10 µM for shoots, whereas in roots hormetic effect was found at 2.5 µM.
Figure 3.5: Straight and inverted U-shaped hermetic curves.
Ha
rmfu
lB
en
efits
Hormetic region
Hormetic region
Ha
rmfu
lB
en
efits
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3.10 CONCLUSIONS
Lead is a highly toxic metal without any known essential role in biochemical functions of plants. Lead has numerous industrial uses and is ubiquitous in the environments. Lead is discharged into the environment both by natural and anthropogenic sources. Lead has very low mobility and solubility in soil and generally exists as adsorbed onto soil particles and/or precipitated in soil solution. Lead has very low soil-plant transfer index. Root absorption is the main pathway of Pb entrance into plants. Lead enters plants mainly via calcium channels. Inside plants, Pb accumulates mainly in root cells with very low root-shoot transfer factor except for hyperaccumulator plants. Lead toxicity interferes with various metabolic processes. It disturbs cell membrane and functionality of different enzymes. It also induces production of ROS and causes DNA damage and lipid peroxidation. Lead-induced ROS are scavenged by various enzymatic and non-enzymatic antioxidants of plant defence mechanisms.
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CHAPTER-4
BIOGEOCHEMICALBEHAVIOUR OF MERCURY IN SOIL-PLANT SYSTEM
4.1 INTRODUCTION
In soil environment, heavy metals occur naturally from pedogenetic processes during weathering of parent material. Heavy metal level seldom exceeds traces and these concentrations are rarely toxic (Kabata-Pendias 2011). Mercury (Hg) element belongs to the same group as that of Znand Cdin the Periodic Table. It has the atomic weight of 200.6, atomic number 80, melting point −13.6°C, boiling point 357°C and density 13.6 g cm−3 (Bjuhr 2007). Mercury is only one of its kinds among metals that is found in the environment in numerous chemical and physical forms like elemental Hg (Hg0), associated with other ions (HgS, HgCl2), inorganic Hg (Hg2+), organic Hg (CH3-Hg) and mercurous chloride or calomel (Hg2Cl2) (Clarkson et al. 2007). Unlike other transition metals, Hg is the only metal that exists in liquid form at room temperatures. Mercury is used in many technological applications like informatics, production of light bulbs and batteries as it is a good conductor of electricity (Raquel Azevedo and Rodriguez 2012). Elemental mercury may also be found in the atmosphere having residence time from 0.5 to 2 years (Schroe der and Munthe 1998) and can be transported with air, hence considered a global pollutant. The Hg can be transported and deposited to remote places even 1000 km away from the source (Johansson and Tyler 2001). Moreover, Hg can also travel and transformed in different environmental compartments (i.e. plants, dust, soil, sediments and atmospheric aerosol), in particulate solids, in aqueous solution and in gas phase (Charlesworth et al. 2011). The two important properties of Hg (high water solubility and volatile nature) make it a non-point source pollutant (Clarkson and Magos 2006).
Natural combustion, weathering of rocks and volcanic eruptions are the natural sources of Hg environmental emissions (Ullrich et al. 2007a, b). As a result of degassing from the Earth crust and evaporation from water bodies, Hg is emitted naturally (Chen et al. 2016). Anthropogenic sources such as energy production, transport, housing, agriculture, mining, landfill, and illegal waste dumping can generate about half of the atmospheric Hg (Schroeder et al. 1998; Gworek et al. 2013). Additional sources include collection, processing and incineration of Hg containing waste materials like batteries and different kinds of industrial wastes (scrubber sludge).
Natural sources play major role in contributing Hg emissions to global atmospheric Hg budget (Pirrone et al. 2010; Driscoll et al. 2013; Dai et al. 2013). Currently, it is estimated that Hg emissions are about 80-600 tons yr-1 from natural sources (Mason et al. 2012); whereas contribution from anthropogenic process ranges between 1926-2320 tons yr-1 (Pirrone et al. 2010). The global Hg emissions into the atmosphere were recorded to 3000 tons in 2005 (UNEP 2013) during post-industrial period however, burning of fossil fuel resulted in enhanced Hg levels in soils by a factor of 3-10 times (UNEP 2013).
Transboundary air pollution convention suggests that Hg is one of the main contaminants that pollute the places far away from the site of its emission (Chen et al. 2016). In environment, Hg is ubiquitous and often its levels reach to an extent that raises concerns for environmental effects and human health (Scheuhammer et al. 2008). Among the top ten hazardous elements, Hg is ranked as sixth because of its persistence, toxicity and its accumulation in the biosphere (Nascimento and Chartone-Souza 2003). The transformations of inorganic Hg to methyl mercury [CH3Hg] + (a specie of Hg which has potential to bio-accumulate in living organisms) is one of the
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major concerns (Karlsson and Skyllberg 2003). At higher tropic levels, foraging behaviour can thus transfer Hg to whole food web and thereby threaten environmental health (Gabriel 2004). Abandoned Hg mines, gold-mining (Qiu et al. 2006), and processes such as recycled Hg processing or the chlor-alkali industry (Ullrich et al. 2007a, b; Reis et al. 2009) are direct or point sources for Hg contamination. Telmer and Veiga (2009) estimated that approximately 1000 metric tons yr-1 of Hg is emitted into environment from almost 70 countries and about 650 metric tons yr-
1 of this amount is discharged into the hydrosphere (soils, lakes, rivers, tailings) while the remaining 350 metric tons yr-1 is directly emitted into the atmosphere.
4.2 GLOBAL USES OF MERCURY
Mercury (Hg) is one of the most commonly used toxic heavy metal found in tectonosphere, asthenosphere, biosphere, hydrosphere, lithosphere and atmosphere. Most of the Hg found in the environment is organo-mercuric and inorganic mercuric salts except in atmosphere. The inorganic mercuric salts such as HgS, Hg(OH)2 and HgCl2 are the wide-spread forms occurring in the environment while, CH3HgCl and CH3HgOH together with other organo-mercurics (phenyl mercury and dimethyl mercury) are main organic forms found in the environment (Zhang and Wong 2007). Mercury has been utilized in a wide range of products over years due to its unique chemical properties such as response to changing temperature and its ability to maintain its volume under varying atmospheric pressure. These properties make it useful in devices designed to measure temperature and pressures. There are numerous Hg containing components that performs concrete function in electrical and electronic devices. Many of these appliances contain ≥1 g of liquid Hg. Moreover, Hg has many industrial uses, including manufacture of dental amalgams, caustic soda, manometers, chlorine, mirror production and gold and silver mining (Chen et al. 2016).
Figure 4.1: Annual world refinery production of mercury (USGS 2016).
Many industrialized nations have banned the utilization of Hg in the manufacturing of industrial products and home appliances. There has been a general shift away from Hg based products (thermometers, batteries) in developed countries, where there is often a higher level of awareness of the environmental and health benefits of shifting from Hg-free products and
1.00
1.25
1.50
1.75
2.00
2.25
2.50
10
3T
ho
usan
ds m
etr
ic t
on
s
Mercury
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processes. Mercury is often put into beauty products like creams soaps that promise to lighten the skin as a common pharmacological compound in certain parts of the world (UNEP 2000). The mine production of Hg increased by 37 % from 148x104 in 2006 to 234x104 tons during 2015 (Fig. 4.1).
4.3 MERCURY LEVELS AND SOURCES IN SOIL
Mercury exists in three oxidation states in the environment like Hg0 (metallic, 0), Hg+1
(mercurous, +1) and Hg+2 (mercuric, +2). Mercury is rare earth element and its concentration in the Earth crust ranges from 0.05-0.10 mg kg-1, the majority of which occurs as the mineral cinnabar (Chen and Yang 2012). Different authors reported different background concentration of Hg in Earth crust and surface soil. For example, average concentration of Hg in soil is reported as 0.043 mg kg-1 (Eriksson 2001), 0.053 mg kg-1 (Licht 2005), 0.061 mg kg-1 (FOREGS 2005), 0.09 mg kg-1 (Burt et al. 2003), and 0.1 mg kg-1 (Kabata-Pendias 2011).
Mining is the main source of Hg in the environment. A number of studies reported variable levels of Hg near mining areas:0.21-3.4 ppm near ancient mining area of England (Davies 1976); 0.2-1.9 ppm in Hg mining of Canada (Kabata-Pendias and Pendias 2011); 5-1710 ppm in Almadén mine of Spain (Millán et al. 2006); 2.6 to 2.9 ppm in Hg mining of France (Kabata-Pendias and Pendias 2011); 12.1-100 ppm in Türkönü Hg mine of Turkey (Gemici and Tarcan. 2007); 8.4-610 ppm in Guizhou-China (Tuzen et al. 2009); 0.09 to 0.22 ppm in Hg mining of Brazil (Kabata-Pendias and Pendias 2011), and 0.079-62 ppm in Aktepe and Gözeçukuru area Turkey (Sasmaz et al. 2015). The variation in Hg soil levels seems due to variation in Hg contents of rocks from which these soils are formed (Table 4.1). Occurrence of Hg in soils significantly correlates with the existence of other heavy metals. For example, Sasmaz et al. (2015) showed a linear correlation (r=0.48-0.54) between Hg and As, Ba, Pb, TI and Sb in Aktepe and Gözeçukuru area of Turkey.
It has been estimated that about 3.2–30×106 kg yr−1 Hg is introduced by the anthropogenic activities whereas, almost 3×106 kg year−1Hg is added by the natural sources globally (Issaro et al. 2009). Combining the both anthropogenic and natural inputs, approximately 806×106 kg of Hg has been added into soils, 741 × 106 kg into the atmosphere and118 × 106 kg into water (Issaro et al. 2009). Current estimates show that annual re-emission of Hg ranges from 4000–6300 t yr-1 (Mason et al. 2012). Majority of the re-emitted Hg is ultimately accumulating in surface soils. Natural Hg emission processes include (i) forest fires (ii) emissions from Hg-contaminated aquatic and terrestrial systems, (iii) emissions from Hg-mineral deposits and (iv) emissions from the volcanoes (Camargo 2002). About 590-930 metric tons of Hg is emitted from forest fires (Brunke et al. 2001). The anthropogenic sources of Hg result from (i) incineration of solid waste (ii) combustion of coal and oil, (iii) pyrometallurgical processes (zinc, lead and iron), and (iv) Hg and gold production (Pai et al. 2000). In comparison to anthropogenic sources, natural and re-emitted Hg emissions act as nonpoint sources so its wide range distributions make it difficult to accurately evaluate emissions and to develop effective control measures.
In the atmosphere, majority of the Hg (80%) is in Hg0 gaseous form (Lindqvist and Rodlhe 1985), which can persist in the atmosphere for > one year because of its high volatility (Slemr and Langer 1992), which causes long distance atmospheric transport of Hg0. Mercury returns to the Earth surface through wet and dry deposits after being transported within the atmosphere. In this way, > 90% of the emitted Hg could accumulate in terrestrial areas (Lindqvist 1991). In North America, 40% of Hg emission has been recorded through discarded fluorescent lamps, batteries and thermometers. Use of barometers in air fields, air ports, weather stations, engine manufacturing, wind tunnels, and off shore installations on ships also contributes Hg release into the environment (Hutchison and Atwood 2003). Mercury pollution also originates from agricultural applications such as fertilizers, pesticides, sewage sludge which discharge Hg into irrigation waters and into the agricultural systems (Hseu et al. 2010). Even though large differences exist between regions, use of Hg in cement production is another main source of Hg, which has enhanced by almost 30% from 2005 to 2009 (US Geological Survey, USGS 2012).
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Table 4.1: Mercury contents in different soil orders and rock types.
Soils order Concentration (mg kg-1) Reference
Podzols 0.008–0.7
Kabata-Pendias 2011
Cambisols 0.01–1.1
Calcisols/Leptisols 0.01–0.5
Kastanozems/Chernozems 0.02-0.53
Histosols 0.04-1.11
Soil Layer/ Rock type Concentration (µg kg-1) Reference
Normal soil 20-150 Mihaljevic 1999
Surface soil 400 Kabata-Pendias 2011
Sedimentary rocks 400 Wedepohl 1978
Carbonate rocks 40 Wedepohl 1978
Sandstone 30 Wedepohl 1978
Crustal rocks 80 Greenwood & Earnshaw 1984
Erosion of the primary mine wastes is a secondary pathway for Hg release into the environment. For example, after small-scale refining of the metal, soluble Hg can leach from calcines (tailings) residues (Li et al. 2008). Soils from Hg mines conventionally contain higher concentrations of Hg that can be readily methylated by biotic reactions once discharged into the free environment especially by the obligate anaerobic sulfate-reducing bacteria (Galloway and Branfireun 2004). In paddy fields, elevated methylated-Hg levels have been reported that were contaminated by adjacent mining areas. In paddy fields near the Wanshan and Wuchuan Hg mines of China, methylated-Hg concentrations in soils were 23 and 20 µg kg−1, respectively (Wang et al. 2012).
4.4 PHYTOAVAILABILITY OF MERCURY
Mercury is absorbed by plant roots via transpiration stream from soil solution. Due to low solubility, Hg availability to plants is low (Baya and Van Heyst 2010). Major portion of absorbed Hg is stored in root cells (Patra and Sharma 2000). However, the highly toxic species of Hg, methylated-Hg, can transfer and translocate from soils to shoots (Bishop et al. 1998). Metals can be up taken from atmosphere via foliar transfer (Shahid et al. 2017b; Xiong et al. 2017). Foliar absorption Hg from the atmosphere is also reported which may occur directly through stomata in leaves (Millhollen et al. 2006). Soil to plant transfer factor can explain Hg behaviour in soil-plant system. Mercury concentration, speciation, availability and plant species are important factors
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affecting transfer of Hg in plants. It has been verified that non-cinnabar Hg forms have high phytoavailability to living organisms such as earthworms (Rumayor et. al. 2015).
Biotic and abiotic processes are responsible to produce elemental Hg in soils. It mostly originates in soil from microbial activities in the A horizon and presence of reductants transform various Hg species (Schluter 2000). When Hg is released through natural processes in soils, precipitation may occur (Song and Van Heyst 2005). Sigler and Lee (2006) suggested that increase in soil temperature caused desorption of bound Hg into gaseous Hg(0) in soils air spaces as a result of which it moves to upper soil layers. The mechanism of Hg(0) formation in soil is not clearly understood, but studies showed that Hg(0) was formed from available Hg(II) by abiotic processes or the desorption of Hg(0) from soil constituents (Gu et al. 2011). The Hg(0) can eagerly adsorb onto the particle surfaces and stay there. Soil is a big pool for Hg in which moisture and temperature alterations affect its speciation.
4.5 SOIL PROPERTIES AND HG BIO-AVAILABILITY
Soil composition can greatly impact Hg biogeochemistry in soils. Compared to water bodies and other biomes, Hg has high persistency in soils (Tangahu et al. 2011). In Histosols and Cambisols, higher Hg levels have been found (Kabata-Pendias 2011). Even though soil has a natural ability to adsorb metal by diverse mechanisms but levels of metals exceeding the soil capacity results to its presence in soil solution (bioavailable form). From all over the world, the average background levels of Hg in various soils range between 0.58-1.8 mg kg-1, and the mean value is 1.1 mg kg-1. Among different soil components, clay can bind/hold high levels of Hg because of its comparatively high surface area (Zhong and Wang 2008). It is well-established that soil clay minerals play a key role in controlling metal–solid partitioning (Kongchum et al. 2011) and play key part in Hg biogeochemistry.
Chloride (Cl-1) is one of the best Hg complexing agents. It has been demonstrated that presence of higher Cl− concentrations limits the Hg(0) presence. Furthermore, decreased organic bound fraction of Hg is also observed in the presence of Cl- (Reddy and Aiken 2001). Experiments showed that high Cl− concentrations caused a shift in Hg speciation mainly from Hg-fulvic acid complexes to Hg–Cl complexes (Reddy and Aiken 2001). The important inorganic sorbents for Hg include hydroxides of Fe, Mn and Al as well as amorphous oxides and clay minerals (Liao et al. 2009). The strong Hg/OH covalent bond is very important for the active binding of Hg with hydroxyl complexes (Schuster 1991). Mercury immobilization may occur due to the formation of apatites, carbonates and Mn and Fe hydroxides/oxides (Oliva et al. 2012). In the context of atomic radius, Hg fits well into carbonates and apatites, while hydroxides/oxides of Mn and Fe are capable to incorporate Hg complexes such as those with OH− or Cl− (Serrano et al. 2012).
4.5.1 Soil pH and Mercury Phytoavailability
Soil reaction (pH) is an important factor affecting the Hg phytoavailability in soil. The biological, chemical and physical properties of soil are influenced by pH as acid-base equilibria are involved in these processes. The solubilisation of nutrients or other minerals in soil solution is dependent on phytoavailability of plant nutrients and pH. The pH may alter the rate of soil organic matter decomposition, activity of microorganisms and consequently the release of different ions. The mobility and availability of heavy metals (including Hg) in soil is affected by soil pH, not only due to its effect on soil characteristics but also due to metal speciation in soil solution (Barrow and Cox 1992). In neutral to alkaline soils, mineral components affect Hg solubility more comprehensively than that in acidic soils (Kwaansa-Ansah et al. 2012). A very small concentration of Hg forms complexes at neutral or at higher pH and extent of adsorption depends upon the interaction among other Hg forms with soil surface and Hg complexation by dissolved organic matter. The extent of complexation of Hg by dissolved organic matter is dependent on its adsorption capacity. Mercury is more accessible in soils with low pH (Gabriel and Williamson 2004). With a decrease in solution pH, Hg ions bind to organic matter more strongly (Schindler et
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al. 1980). The pH differently influences Hg mobilization in different soil environments as it fluctuates in acidic to alkaline environment. The minimum extent of Hg dissolution was found around pH 3 (Xu et al. 2014). Soil pH highly influences the impact of polysulfides on Hg solubility. By using a conceptual wetland soil/sediment, chemical speciation of Hg and [CH3Hg] + was modelled at pH 4.0 and 7.0. In this model, typical concentrations of thiols, sulfides, Hg and [CH3Hg] + were used. Thiols were out-competed by Hg–polysulfide complexes at pH 7. The Competition ability of polysulfides was the least when pH decreased to 4 and at this pH, 50% of Hg binding to thiols and 50% to polysulfides and bisulfides werereported (Skyllberg 2008). Mercury adsorption increases from pH 2 to 4, with maximum adsorption between pH 4 and 5 (Yin et al. 1996). The magnitude of Hg desorption fluctuates with change in soil pH that may be due to strong adsorption of metal hydroxide complexes, competition of protons for binding sites, hydrolysis of Al on exchange sites, and acid catalysed dissolution of reactive oxide sites.
Unlike other heavy metals (Zn and Cd), the effect of pH is quite different on ionic balance of Hg species (Appel and Ma 2002). As pH rises, hydroxide of Hg is formed, i.e. Hg(OH)2,
HgOHCl and HgOH+ (Yin et al. 1996). The pH effect is also influenced by the concentration of carbon nanoparticles. At high mass ratios of carbon to Hg, multi-walled carbon nanotubes induce a slight effect on Hg sorption between pH 5 and 7 (Tawabini et al. 2010). In organic layer compared to mineral top soil layer, higher total Hg concentrations but lower pH was observed, certainly indicating the crucial role of pH in the accumulation of Hg in soils. However, affinity of Hg to binding sites in organic matter decreases with increasing soil pH resulting in release of Hg into soil solution (Zhou et al. 2015).
4.5.2 Soil Organic Matter and Mercury Phytoavailability
Organic matter is the leading factor in controlling Hg movement in soils (due to its high affinity or Hg (Kwaansa-Ansah et al. 2012; Zhong and Wang 2006). The presence of organic matter affects phytoavailability and fractions of Hg in soils. Natural organic matter affects Hg mobility, solubility, speciation and toxicity. A number of organic compounds or humic substances have higher affinity for Hg owing to the presence of functional groups, such as aromatic, carboxylic, hydroxylic and particularly the S-containing ligands (Kabata-Pendias 2011). Moreover, presence of sulfhydryl and thiophenolic groups in humic substances binds Hg(II) strongly (Xia et al. 1999). Skyllberg et al. (2005) reported that methyl-Hg was found dominantly to form complexes with thiol groups in soil over a wide range of pH. Various experiments have indicated that the presence of humic acid increased Hg adsorption in Fe rich soils (Lumsdon et al. 1995). Hence, Hg leaching from soils to aquatic system decreases due to presence of humic substances in soils. At the same time in tropical soils, both the reduction and complexation of Hg(II) to volatile Hg by humic substances can occur (Rocha et al. 2000).
It is described that the quantity and the type of organic substances regulate the partitioning between aqueous and solid phases that plays an important role in the circulation of Hg in soil profiles (Semu et al. 1987). Distribution of Hg varies with soil depth (Manceau et al. 2015), depending on organic matter contents, microbial community structure and mineralogical composition (Hansel et al. 2008). The biogeochemical characteristics of soils may have control on Hg dynamics like its mobilization and transformation in soil (Poulin et al. 2016). In soils, Hg(II)-organic matter speciation dominates by forming complexes with thiol group (Nagy et al. 2011). Likewise, the co-mobilization of Hg with DOC (Haitzer et al. 2002) from soils to streams is due to complexation of Hg(II) with thiol group in DOC of soil (Dittman et al. 2010). The Hg(II) can also be reduced to Hg0 through abiotic [e.g., aqueous Fe(II), Fe(II)-bearing minerals, redox-active DOC groups] (Bone et al. 2014) and biotic pathways (dissimilatory metal-reducing bacteria) (Hu et al. 2013) that ultimately takes its way to water bodies.
4.5.3 Soil Microbial Activity and Mercury Behaviour
In addition to soil chemical and physical processes, Hg(0) production in soils is influenced by microbial activity (Choi and Holsen 2009; Leveque et al. 2014). Mercury-specific detoxification pathways are used in bacterial transformations of toxic Hg to less toxic Hg species (Barkay et al.
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1991) or microbial detoxification of reactive oxygen species (ROS) is related with non-specific microbial reduction of the Hg(II) (Siciliano et al. 2002). On the other hand, anaerobic micro-organisms and sulfate reducing bacteria convert Hg(II) to MeHg in soils (Podar et al. 2015) which was even a more toxic form of Hg (Ullrich et al. 2001). Bioavailable fraction of Hg(II) in soils is more important for the production of Hg(0) rather than the total Hg(II) contents. Organic matter contents change the affinity between rock deposits and Hg(0) under anoxic conditions. It also enhances the rate of reduction of Hg(II) to Hg(0) by soil biota (Bouffard and Amyot 2009). So, it is possible that microbial reduction may stimulate Hg(0) production in soils. Many studies have investigated the role of organic matter in methylation of Hg to methylmercury (MeHg) by bacteria (Fitzgerald et al. 2007). Zhong and Wang (2009) studied the relative importance of bacterial sorption in controlling Hg partitioning and bio-availability in sediments. They showed that numerous types of yeasts and bacteria have been reported to alter the reduction of cationic Hg(II) to elemental state Hg(0). Several microorganisms such as Bacillus and Streptomyces can also mediate elemental Hg to its cationic forms through oxidation.
High MeHg bio-magnification has been observed in terrestrial food web which is a rising concern about Hg bioaccumulation (Newman et al. 2011). Earth worms are efficient in accumulating Hg and their bio-accumulation factor of MeHg ranged from 10 to 249 (Rieder et al. 2011) compared to < 10 for inorganic Hg (Burton et al. 2006). In environmental risk assessment, bio-accumulation indicates a basic and critical role. Mechanistically from all possible pathways, earthworms could accumulate heavy metals including pore-water and diet, e.g. soil ingestion (Kaschak et al. 2014). To accomplish the requirements of energy and nutrition, earthworms ingest soil at high rates (Curry and Schmidt 2007). Several microbial and non-microbial processes generally enhance Hg vaporization. Moreover, volatilization rate of Hg is generally the highest in soils having intense microbial processes (Andersson 1979). Due to gene transfer mechanism, bacteria are adapted to withstand high levels of Hg in flood plain soils (Oregaard and Sorensen 2007). Clays also could complex on bacterial cell wall surface as a process of Hg resistance (Tazaki and Asada 2007).
4.6 SOIL-PLANT TRANSFER OF MERCURY
4.6.1 Molecular Understanding of Mercury Absorption by Plants
Generally, plants play a comparatively inert role in the biogeochemistry of Hg compounds in soils. So far, no naturally occurring plant species has been identified which can accumulate and remove Hg from soil. The mechanisms of Hg absorption and transport have not been studied at extensive level as for other heavy metals. Mercury absorption in plant takes place via the same route (roots) as for another essential micronutrient (Patra and Sharma 2000). Mercury absorption and accumulation in different plant parts have been studied in several plant species such as Verbascum Thapsus, Anchusa arvensis, Onosma, Alyssum saxatile, Glaucium flavum, Centaurea cyanus, Silene compacta and Phlomis (Sasmaz et al. 2015), Beta vulgaris, Triticum aestivum, Trifolium repens, Brassica napusand Salix babylonica (Greger et al. 2005), Lavandula stoechas (Sierra et al. 2009), Jatropha curcas (Marrugo-Negrete et al. 2016) and wild plant species (Moreno-Jimenez et al. 2006). Plants can absorb Hg from soils or nutrient solution but, generally, translocation to shoots is very small. Lupinus albus and Brassica napus showed low affinity active component on Hg influx with the Km 1.12 × 10-13, 1.87 × 10-13 M and Vmax 659, 1645 nmol Hg FW−1 h−1, respectively (Esteban et al. 2013). On the other hand, several bacteria accumulated inorganic Hg in their bodies. For phyto-accumulation purposes, instead of accelerating prevailing processes in plants, a different technique was anticipated with completely new pathway by encoding two bacterial genes in plants that together transformed MeHg to volatile elemental Hg (Ruiz and Daniell 2009). Organo-mercurial lyase is encoded by MerB, which converts MeHg to Hg[II]; mercuric reductase is encoded by MerA, which causes reduction of Hg[II] to Hg[0] (Fig.4.2) (Summers 1986). Transgenic MerA Arabidopsis thaliana plants showed considerably higher volatilized elemental Hg and tolerance to Hg[II] (Rugh et al. 1996). Similarly, transgenic MerB Arabidopsis thaliana plant species showed high tolerance to MeHg and other organo-mercurials (Bizily et al. 1999). Infact, MerB plants are capable to convert toxic MeHg to less toxic ionic Hg.
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The MerA-MerB double-transgenic has been shown to be more tolerant to organic Hg compound as compared to their MerA, MerB and wild-type parents (Bizily et al. 1999). While MerB encoded plants were almost 10-times more tolerant to organic Hg compared to wild-type plants.
Figure 4.2: Mercury absorption and transformation in plants.
4.6.2 Competition between Mercury and Cations for Absorption
Competitive interactions at absorption sites between toxic and essential ions take place depending on the ionic size and charge. By examining and comparing numerous essential nutrients for their chemical properties, Hg has the closest chemical relatives like Zn, Cu and Fe, so that their transporters may also facilitate Hg influx into plant roots. Divalent metal ion transporter 1 (DMT1) is a divalent transition metal/H+ co-transporter from the natural resistance-associated macrophage (NRAMP) family. The DMT1 is reported to transport a number of essential divalent metal cations, such as Zn, Ni, Fe, Co and Mn (Zalups and Koropatnick 2010). It has been suggested that DMT1 was involved in the absorption of Hg[II] (Vázquez et al. 2015), Cd2+ and Pb2+ (Bressler et al. 2004). Cailliatte et al. (2010) reported that NRAMP1 acted as a Mn transporter in Arabidopsis thaliana may also be involved in Hg[II] transport in crop plants such as Brassica napus.
Metal transporters belonging to the ZIP(ZRT, IRT-like protein) family can also mediate the transport of numerous trace elements, including Fe, Cd, Co, Zn and Mn (Pedas et al. 2008), suggesting that those may also transport Hg[II]. Mercury synergism with Ni and antagonism with Cu at absorption stage suggest that its absorption by plant roots occurs through one or many essential nutrient transporters. Calcium channels or aquaporins blocked by Hg (Johansson et al. 2000) could also partially participate in the Hg absorption by Lupinus albus. Some studies also have demonstrated that selenium (Se) decreased Hg toxicity in living systems (Belzile et al. 2006). Mounicou et al. (2006) demonstrated the Se–Hg antagonism in plants. Presence of Se can significantly decrease the absorption and accumulation of Hg in plant cells, which could be due to the formation of Hg–Se- complex in plant roots, and Wang et al. (2014) showed that this complex cannot be transferred to aerial tissues of plants.
Hg(II)
DMT
ZIPHg(II) MeHg
MerA
Hg-PC
Intercellular methylation
Hg(0)
Complexation and detoxification of Hg
As(III)-PC
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4.6.3 Mercury Sequestration into Plant Roots
Many reports have revealed that plant roots accumulated Hg when they were cultivated in Hg-contaminated soils. Lab-scale studies have shown that plant roots absorbed Hg from the growth medium and roots accumulated higher Hg than that by shoots (Cavallini et al. 1999). There are substantial evidences signifying that root to shoot translocation of Hg is very low in wild type plants (Suszcynsky and Shann 1995). Heaton et al. (2005) revealed that Hg mainly accumulated in Oryza sativa roots with a very little translocation to shoot tissues when grown in soils and hydroponics. Plants detoxify Hg by complexing it with phytochelatins (PCs) and subsequently sequestering it into vacuoles. Two ABCC-type transporters, AtABCC1 and AtABCC2, detoxify Hg[II] and plant mutants with these genes (atabcc1 single or atabcc1 atabcc2 double) displayed a hypersensitive phenotype against Hg[II]. This complexation and sequestration of PC-Hg complex is responsible for high Hg accumulation in plant roots. However, some authors found that PCs could only sequester Hg[II]) but not MeHg (Krupp et al. 2009). They also showed that Oryza sativa plants treated with Hg[II]) (10 mg Hg L-1) contained as much as 253 mg kg-1 in comparison to CH3Hg fed plants with 121 mg kg-1 in roots. Godbold and Hűtterman (1988) suggested another mechanism by which plants retained significant amount of Hg into roots by complexing Hg to cell walls.
4.6.4 Mercury Translocation to Plant Shoots
Mercury has very low phytoavailability, because of its preferred adsorption onto soil particles or precipitation in solution (Lucena et al. 1993). After absorption, Hg is primarily sequestered in plant roots (Patra and Sharma 2000). On the other hand, small amount of Hg reaches shoots by translocation. It is reported that total Hg concentration in soil solution did not affect Hg concentration in both the roots and shoots of plants. This shows that Hg level in soil solution has no effect on its translocation to plant shoot tissues. This is because of the fact that Hg occurs in various organic and inorganic forms, which vary in phyto-availability (Ravichandran 2004). Moreover, volatilization feature of Hg can also be the reason of week relationship between Hg concentration in soil solution and plant shoots. It is well-known that Hg volatilizes from plant leaves surface to atmosphere (Leonard et al. 1998), which may contribute to poor correlation between Hg level in soil solution and roots or shoots. Soil-plant transfer of Hg may follow Langmuir-type correlation, which shows that plants have low absorption when Hg in soils is at high levels. However, when Hg level in soil is low, plant Hg absorption increased linearly. On the other hand, at high applied doses of Hg, absorption of Hg by plants increased more gradually.
Some transgenic plant species can accumulate high level of Hg in their shoot tissues. For example, Sasmaz et al. (2015) reported that eight plant species (Verbascum Thapsus, Anchusa arvensis, Onosma, Alyssum saxatile, Glaucium flavum, Centaurea cyanus, Silene compacta and Phlomis) were capable to translocate significant amount of Hg (almost 50% of the absorbed) towards their shoot tissues. Various plant species such as Oryza sativa (Heaton et al. 2003), Nicotiana tabacum (Heaton et al. 2005), Arabidopsis thaliana (Rugh et al. 1996), Populus deltoides (Che et al. 2003) and Liriodendron tulipifera (Rugh et al. 1998), overexpressing merA were able to accumulate and tolerate ten-times higher Hg[II] concentrations compared to non-transgenic plants. These transgenic plants also exhibited higher volatilization of Hg[0] compared to controls. Transgenic plants expressing high merA are reported to grow well compared to non-transgenic plants on Hg-contaminated soil (Heaton et al. 1998). Methyl-mercury is more toxic than metallic or ionic Hg and has relatively high potential of biomagnification in the food chain. In this way, MeHg can induce serious hazard to human and animals. It is reported that merA can only convert Hg[II] to Hg[0] and was unable to defend plants against MeHg toxicity. In order to protect plants from MeHg toxicity, both the merA and merB genes are required. Transgenic Arabidopsis thaliana plants expressing merB showed more tolerance to MeHg compared to non-transformed plant species (Bizily et al. 1999). These results confirmed that expression of merB can mediate MeHg tolerance of plants.
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4.6.5 Mercury Accumulation in Edible Plant Parts
Environmental pollution by Hg is a serious environmental concern. Its concentration may exceed the permissible limit in some edible plants and mushroom species (Falandysz and Brzostowski 2007). According to joint FAO/WHO Expert Committee on Food Additives (JECFA), maximum weekly intake for MeHg is 1.6 µg kg-1 body weight and 4 µg kg-1 body weight for inorganic Hg. Plant-derived food plays minor role in Hg accumulation and exposure in human food chain, except when cultivated near to Hg mining areas (Rodrigues et al. 2012; Rothenberg et al. 2014). Majority of MeHg, which is the most toxic and bioavailable form of Hg, enters into human food chain by consuming fish and other sea foods (Mergler et al. 2007). Organic forms of Hg are of greater concern as the absorption rate of MeHg by the human body is about 95% (WHO 1990). However, recent experiments have demonstrated that Oryza sativa consumption can act as a main pathway of MeHg exposure to humans in Hg mining areas (Zhang et al. 2010). Generally, Hg concentration in human food items is below 20 μg kg-1, predominantly present in inorganic forms (WHO 1991). Higher concentrations of MeHg (140 μg kg-1) were recorded in Oryza sativa grown in Wanshan Hg mining site, China (Horvat et al. 2003). Similarly, Qiu et al. (2008) stated that Oryza sativa grown near Hg mining areas accumulated >100 μg kg-1 MeHg in grains, that is 10 to 100 fold higher than other crops (Zea mays, Brassica rapa, Nicotiana tabacum, and Brassica oleracea) grown in the same areas. Shariatpanahi and Anderson (1986) concluded from the analysis of different vegetables (ranging from 0.06 in Allium cepa to 0.13 µg g-1 in Artemisia dracunculus) irrigated with waste water contained appreciable amounts of Hg in edible plant parts. Although these levels may not produce acute toxicity in adults but repeated exposure with other environmental sources may induce chronic effects in children. Cynadon dactylon grown in soils containing 50 mg Hg kg-1 accumulated 80-800, 0.4-1.2, and 1.4-37.6 mg Hg kg-1 in roots, stems and leaves, respectively, making them potentially dangerous to grazing animals (Weaver et al. 1984). Open top chamber experiment conducted to explore the relationships between air Hg concentration and leaves of Lactuca sativa, Raphanus sativus, Medicago sativa and Lolium perenne showed a positive correlation between air Hg concentration and accumulation in vegetables and grasses (Niu et al. 2013). Under low and high-Hg treatments, Hg in vegetables grown on Hg contaminated soils showed Hg concentration averaging to 20x10-4 and 35x10-4 mg kg−1, respectively (Ding et al. 2014).
Some species of mushrooms accumulate high levels of Hg in their fruiting bodies. Melgar et al. (2009) analysed 238 samples from 28 wild growing edible mushrooms species from the province of Lugo (NW Spain). The mean Hg concentration was found in Boletus pinophilus (4.5 mg kg-1 DW), Agaricus macrosporus (3.7 mg kg-1 DW, Lepista nuda (3.1 mg kg-1 DW) and Boletus aereus (3.3 mg kg-1 DW), while the lowest was found in Agrocybe cylindrica (0.26 mg kg-1 DW) and Fistulina hepatica (0.22 mg kg-1 DW) in fruiting bodies of mushrooms. From these results, author concluded that species Agaricus macrosporus, Boletus pinophilus, Boletus aereusand Nepeta nudas should be consumed in low amounts to avoid Hg risk.
4.7 EFFECTS OF MERCURY ON PLANTS
At the cellular level, the blocking of important molecules (polynucleotides and enzymes), substitution of essential metal ions from molecules, interruption in the transport of essential ions, disruption of cell organelles and membranes and denaturing of proteins are the possible mechanisms by which heavy metals induce toxicity to plant tissues (Ann. Table 3) (Azevedo and Rodriguez 2012). Hence exposure to high levels of Hg could induce negative effects on plant growth and metabolism (Patra et al. 2004). Higher concentrations of Hg make it strongly phytotoxic to plant cells causing toxicity and physiological disorders (Zhou et al. 2007). The relative toxicity of different metals to plants may be different for different plants depending upon the genotype and the experimental conditions. Plants (especially agricultural crops) absorb excessive amounts of metals like Hg and transfer them into food chain (Devkota and Schmidt 2000).
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Accumulation of Hg in plants impedes various cellular-level metabolic functions, causing reduction in development and growth, but the exact mechanism is not clear yet (Kabata-Pendias 2011). The possible mechanism of Hg phytotoxicity may be its high affinity to bind and react with the sulfhydryl groups and phosphate groups, change in cell membranes permeability, essential ions replacement and disruption in functioning of critical or non-protected proteins (Patra et al. 2004). Both inorganic and organic Hg forms have been shown to cause accumulation of iron and deficiency ofMn, Mgand K(Boening 2000). Some major toxic effects of Hg in plants include increased lipid per-oxidation (Cho and Park 2000) decreased photosynthesis, increased transpiration, low water uptake and synthesis of chlorophyll (Godbold and Huttermann 1986). Mercury absorption, toxicity, and compartmentation in different tissues are different between shoots exposure to elemental Hg0 and roots exposure to Hg2+. Mercury accumulation is reported in shoot tissues with no transfer towards roots when they were exposed to foliar Hg0, while root-exposed plants accumulated Hg with its transfer to shoots. A level ≥ 1.0 µg mL-1 decreased root and shoot growth even with low toxicity to tissues at higher applied doses (Suszcynsky and Shann 1995). Mercuric cations can disturb many metabolic functions where non-protected or critical proteins are involved due to its high affinity for sulfhydryl groups.
Mercury ion has ability to bind at two locations of a protein molecule without disturbing the chain, and it can also bind with two neighbouring chains at a time. Mercury may cause protein precipitation when it is present in sufficiently high concentration. As organomercurials, Hg atom still has a free electron, so a monovalent ion is formed by such salts. The mercurials may decrease activities of numerous intracellular enzymes non-specifically, and may also cause in vitro inhibition of numerous specific thiol-containing respiratory enzymes. It disturbs both dark and light reactions of photosynthesis. It causes a strong inhibition of electron transport chain during photosynthesis; the most sensitive target is photosystem II. The Hg[II] ions interact with Mn clusters which is present in the oxygen evolving complex and it is present on the donor side of PSII. It also interacts with the intermediate Zn ions which are located in D1 and D2 proteins. Mercuric chloride causes an oxidation of PSI (P700) in the dark while donor side of PSII is affected by inhibiting chloride binding and/or function. The Hg[II] ions bind with amino acids of chloroplast proteins to form organo-metallic complexes and thus a polypeptide (33 kDa) of PSII sub-membrane is depleted (You et al. 1999). A decrease in the amount of chlorophyll is reported by Hg treatment and it resulted in breakdown of thylakoid. Moreover, under Hg stress, the activity of nicotinamide adenine dinucleotide phosphate (NADPH): protochlorophyllide oxidoreductase (POR) is inhibited, which are responsible for the biosynthesis of chlorophyll (Lenti et al. 2002). Inhibition of water channels in the membranes of higher plant cells was also reported under Hg stress. In Brassica napus seedlings, Hg (at concentration >1 mg L-1) was found to accelerate lipid per-oxidation of membranes, increase the membrane permeability and interrupts membrane structural integrity (Ma 1998). The hydraulic conductivity of Triticum aestivum root cells was decreased by mercuric chloride (Zhang and Tyerman 1999). In leaf discs of Quercus robur, HgCl3 (120 µM) induced a concentration-dependent decrease in both GSH levels and non-protein thiol and initiation of a time- and dose-dependent glutathione S-transferase (Gullner et al. 1998).
4.7.1 Mercury Toxicity to Plant Growth
Seed injuries and decreasing seed viability are related to many forms of Hg exposure. When Hg form S–Hg–S bridge by interacting with SH groups, the stability of the group is disrupted, as a result embryo growth and seed germination is affected (tissues rich in SH ligands). A decrease in overall seed germination was reported in some Cinnamomum cassia species upon Hg and Pb exposure at two different levels (10 mg L-1, 50 mg L-1) (Khan 2013). In Arachis hypogaea, growth parameters like seed germination, root and shoot lengths decreased due to mercuric chloride solutions (1, 2, 3, 4, and 5 mg L-1) as compared with that for control treatments. A significant decrease in seed germination was recorded at higher concentrations of Hg (3 mg L-1, 4 mg L-1 and 5 mg L-1). At exposure of 5 mg L-1 mercuric chloride, seedling growth was significantly affected compared to that with the control. Root growth and shoot growth of Arachis hypogaea was negatively affected at 3 mg L-1 exposure (Abraham et al. 2012). On the other hand, stimulation in growth of Triticum aestivum seedlings was reported at a low
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concentration of Hg (10 µM). Possible reason for this increased germination can be the increased activities of enzymes like amylase, proteinase and lipase and faster decomposition of endosperm and the respiration rate at lower Hg concentration. Inhibition in primary roots elongation of Zea mays and an inhibition in the gravimetric response of seedlings were shown by HgCl2 (Patra et al. 2004). Furthermore, increasing concentration of Hg2+ in nutrient solution decreased both the root and shoot biomass of Oryza sativa due to its high Hg absorption by seedlings. A Hg concentration of 1.0 and 2.5 mg L−1 resulted in 50% reduction in root biomass while Hg2+ concentrations of around 0.5 mg L−1 resulted in 50% decrease in shoot biomass compared to control (Du et al. 2005). Another study reported a significant decrease in root growth of Oryza sativa seedlings at 12.5 mM Hg in the culture solution compared to the control; while an exposure to 100 mM Hg resulted in complete inhibition of root growth (Chen et al. 2012).
Prolonged Hg exposure decreased levels of photosynthetic pigments viz. carotenoids and chlorophylls (Lomonte et al. 2010). High concentration of Hg (40 μM Hg) inhibited the growth of Sesbania grandiflora cells when grown in culture medium. It might be due to high accumulation of Hg (Israr and Sahi 2006). Concentration and time-dependent decrease in root and shoot lengths of Cucumis sativus seedlings were reported at various applied levels of Hg (0 to 500 µM). This Hg-induced decrease in growth of Cucumis sativus could be due to decrease in pigment contents. It might be related to Hg-induced oxidative stress, which induced lipid per-oxidation (Cargnelutti et al. 2006). In higher plants, Hg can inhibit water absorption through aquaporin on plasma membranes. Iqbal et al. (2014) have reported Hg-induced decreased percentage of seed germination, decreased shoot and root length and damage to seedling dry weight in Albizia lebbeck. A complete inhibition of germination was recorded at 1.5 and 1.7 mM Hg in Cucumis sativus and Triticum aestivum, respectively, while other metals did not present such results even at very high concentration (8.0 Mm). Mercury was responsible for the highest inhibition of root elongation, germination, and hypocotyl and coleoptile growth (Munzuroglu and Geckil 2002). Exposure to high Hg concentration (7 mM) showed maximum decrease in seedling length (70%), seed germination (42%), seedling dry weight (47%) and root length (66%) of Vigna radiata (Iqbal and Siddiqui 2015).
4.7.2 Mercury Genotoxicity
The Hg-induced genotoxicity to plants varies with its applied level, oxidation state and the duration of exposure (Sharma et al. 2012). Exposure to Hg decreased mitotic index in the root tip cells in Allium cepa and enhanced the frequency of chromosomal aberrations, which was positively correlated to the applied level and duration of exposure (Patra and Sharma 2000; Sharma et al. 2012). In relation to genotoxicity, mercuric ions have a tendency to form covalent bonds with DNA (Sharma and Talukder 1989).
The Hg has been reported to induce both mutagenic and clastogenic effects in higher plants. It has been demonstrated that metal caused chromosomal abnormalities and also decreased the rate of cell division. Generally, metal-induced genotoxic effects are more distinctive at higher applied levels for longer duration (Bhowmik et al. 2000). Low levels of Hg can aggravate c-mitosis, chromosomal aberrations, sister chromatid exchanges and spindle alterations (Patra et al. 2004). Mercury genotoxicity affects cell division due to Hg binding with nucleic acids, which cause chromosomal alterations and inhibition of spindle formation in cells (Cavallini et al. 1999). Interaction of Hg ions with thiol group of amino acids and nucleosides is responsible for Hg toxicity. However, some specific effects of Hg such as mutagenesis cannot be attributed to this process. In this case, Hg interaction with nitrogen and oxygen atoms of purine/pyrimidine bases and ribose/ribophosphate groups could be possible action of Hg toxicity. Impairing nucleobase and purine/pyrimidine bases can be due to the ligation of Hg to these bases, which ultimately causes the mutagenicity. By this way, Hg can bring modifications in amino acid orders of proteins to be synthesized (Zalups 2000).
The chromosomes of Allium sativum, Vicia faba and Allium cepa were damaged when exposed to Cd, Pb and/or Hg. They also revealed polyploid, C-karyokinesis, chromosomal bridges, chromosome fragmentation, chromosomal rings, chromosome fusion, micro-nuclei and
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nuclear decomposition upon heavy metal exposure (Duan and Wang 1995). The Hg-induced genes encoding proteins was found to be involved in chlorophyll synthesis, P450- mediated biosynthesis of secondary metabolites and cell wall metabolism (Heidenreich et al. 2001). Mercury exposure resulted in up-regulation of enzymes genes involved in coumarins, caffeoyl, and other monolignols required for lignin biosynthesis (Montero-Palmero et al. 2014). Lignification of cell walls decreased cell wall extensibility (Passardi et al. 2004) and thus may be principal contributing factor in root and shoot growth inhibition. Among the other regulated genes, several stress related proteins such as glutathione, heat shock, pathogenesis related, glutathione/ascorbate cycle have been reported (Montero-Palmero et al. 2014).
4.7.3 Mercury-Induced Oxidative Stress
Mercury stress is responsible for inducing significant generation of ROS (Table 4. 2) (Li et al. 2013). Mercury-induced oxidation damages the cellular functions by membrane damage, lipid peroxidation, and DNA and enzyme inactivation (Cargnelutti et al. 2006; Gill et al. 2012). Scavenging systems that include GSH controls the level of ROS in these conditions (Elbaz et al. 2010; Sahu et al. 2012). Plants experience “oxidative stress” when ROS are not effectively removed. Cells can confer oxidation and variation of cellular DNA, proteins, amino acids and membrane lipids when ROS are produced in excess (Sahu et al. 2012; Shahid et al. 2014a). Oxidative injuries caused by oxidative stress induce decrease in plant development and growth (Cargnelutti et al. 2006). Fenton reaction is induced due to biochemical toxicity of Hg when it reacts with sulfhydryl groups of proteins. As a result, oxidative stress causes lipid peroxidation. Furthermore, over production of ROS decreases chlorophyll contents and biomass production (Patra et al. 2004). The O2
•- is the earlyresponding ROS species under Hg stress. Generation of O2
•-may activate the production of more vigorous ROS like HO• and H2O2, each of which may start peroxidation of macro-molecule such as lipids and proteins (Elbaz et al. 2010; Gill and Tuteja 2010).
Mercury exposure caused an oxidative damage in Suaeda salsa at concentration of 20 μg L−1 (Wu et al. 2011). Disruption of bio-membrane lipids and cellular metabolism is directly correlated with this process (Cargnelutti et al. 2006). The Hg-induced generation of ROS is associated with oxidative damage which is expressed as lipid peroxidation and damage of membrane integrity (Meng et al. 2011). Proteins and membrane lipids are highly susceptible to be attacked by free radicals (Elbaz et al. 2010; Shahid et al. 2015b). These are also considered as a reliable sign of oxidative stress in plants. The most easily described symptom of oxidative stress is lipid peroxidation (Shahid et al. 2012c) and is an indicator of enhanced oxidative damage (Zhou et al. 2009). The peroxidation of lipids possibly starts with the hydroxyl radical. In shoots of Zee mays, lipid peroxidation was evidently induced by Hg and Cd (Campo et al. 2006). Both metals were responsible to cause a low, yet significant increase in lipid peroxidation. Protein oxidation was also boosted by both the Hg and Cd in roots and shoots.
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Table 4.2: Mercury-induced enhanced production of ROS in different plant species grown in
hydroponics
ROS Plant species Exposure level Duration References
O•2-, H2O2
-, OH-
Triticum aestivumφ 0.32, 10.9, 30 mg kg−1
NM Li et al. 2013
O•2-, H2O2
- Oryza sativa 12.5,100 mM 1 day Chen et al. 2012
O•2-, H2O2
-, OH-
Cucumis sativus 0 to 500 µm 10,15 days Cargnelutti et al. 2006
O•2-, H2O2
- Zea mays 6, 30 µm 7 days Rellan-Alvarez et al. 2006
H2O2- Medicago sativa 3, 10,30 μM 1 day
Ortega-Villasante et al. 2007
O•2-, H2O2
- Oryza sativa 0.2 Mm 10 days Wang et al. 2009
H2O2- Medicago sativa 10–40 μM
6, 12, 24, 48, 72 hrs
Zhou et al. 2007
O•2-, H2O2
- Medicago sativa 0–40 μM 7 days Zhou et al. 2008
H2O2- Triticum aestivum 2.5, 5, 10,25 μM 8 hrs Sahu et al. 2012
H2O2- Oryza sativa 10 μM 15 days
Mishra and Choudhuri 1999
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide. Φ Grown in soil.
A sensitive biomarker of oxidative stress to proteins is carbonyl contents. In Cucumis sativus seedlings, an increase in the contents of carbonyl and MDA was recorded which showed a substantial oxidative damage due to high HgCl2 exposure (Cargnelutti et al. 2006). There was a notable carbonyl accumulation in Zee mays exposed to 30 µM Hg, indicating an acute stress oxidation (Alvarez et al. 2006). An Increase in MDA contents after 1h with 25 µM Hg treatment was measured in plant seedlings which confirm lipid peroxidation and oxidative stress under Hg treatment (Chen et al. 2012). The accumulation of O2 and H2O2 in leaves of Medicago sativa in a dose-dependent manner was reported under Hg stress (Zhou et al. 2008). The Hg2+ also triggers the NADH oxidase, which is primarily responsible for the production of H2O2 in plants (Mittler 2002). Increased activity of NADPH oxidase and resultant membrane damages are also reported in Medicago sativa under Hg stress (Zhou et al. 2008).
Enhanced production of ROS is an important indicator of stress caused by heavy metals. It was testified that Hg rapidly induced ROS generation in Oryza sativa roots. Zhou et al. (2007) revealed lipid peroxidation induction in Medicago sativa plants after two days of exposure to Hg. Lipid peroxidation processes in plants are catalysed by enzymes such as lipoxygenase (LOX). Enhanced LOX activity has been reported in Arabidopsis leaves under heavy metal stress including Hg (Skorzynska-polit et al. 2006). Thus, exposure to high levels of Hg may induce
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enhanced ROS production, LOX activity and lipid peroxidation, which may disturb cell redox status in Oryza sativa roots (Chen et al. 2012). Mercury-induced over-production of ROS has been found in Triticum aestivum, Oryza sativa, Cucumis sativus and Zea mays (Li et al. 2013; Chen et al. 2012; Cargnelutti et al. 2006; Sahu et al. 2012).
4.8 MERCURY DETOXIFICATION IN PLANTS
Plants have adapted mechanism for detoxification against metal exposure and accumulation to decrease their harmful effects primarily based on chelation and subcellular compartmentalization. Plants can tolerate metals in three ways: sequester heavy metals in their vacuoles, bind them with various thiol compounds and pump them out at the plasma membrane. Literature has documented the formation of PCs and synthesis of PC-Hg complex in response to Hg stress. Some studies (Elbaz et al. 2010) have reported mechanisms of Hg-induced oxidative stress and detoxification in higher plants. While data are limited for explaining the Hg induced antioxidative mechanism at cellular level. Mercury concentration in the growth medium correlates positively with antioxidants (GSH and NPSH) induction and activities of antioxidative enzymes (Israr and Sahi 2006).
4.8.1 Vacuolar Compartmentalization of Mercury
Heavy metal ions may reach at toxic levels when they are not readily used in cell metabolism. An active metabolic process produces chelating compounds when a metal including Hg exceeds its threshold level inside the cell. Specific peptides (such as PCs and MTs) are used to chelate metals in the cytosol. Specific subcellular compartments serve for the sequestration of heavy metals after their chelating process. Many small molecules like organic acids, amino acids and phosphate derivatives also take part in metal chelation process inside cells (Verbruggen et al. 2009). When intracellular concentration in plants reaches a higher level, plants export Hg ions to the apoplast or are compartmentalized by using their efflux pumps. Vacuoles are the main sink for metal ions including Hg (Verbruggen et al. 2009; Shahid et al. 2014a).
4.8.2 Mercury Chelation by Phytochelatins
Phytochelatins were first isolated in Cd exposed cell suspension cultures (Grill et al. 1985). From that time, PCs have been identified in some eukaryotes, including higher plants. In addition to Cd, plant exposure to some other heavy metals such as Pb, Ni, Cu and Hg also causes the synthesis of PCs. When plants are exposed to Hg, PCs are produced in plant cells from GSH by the activity of constitutive enzyme PC-synthase. Many studies have reported the role of PCs in heavy metal homeostasis and detoxification (Hirata et al. 2005; Dago et al. 2014). Phytochelatins chelate metal ions in the cytosol and then metal-PC complex is transported to vacuole (Yadav 2010). In this way, PCs could protect plants from the negative effect of toxic metals.
Plants generally employ a two-step process to detoxify metal ion toxicity. In first step, metal ions are bound to PCs followed by metal-phytochelatin complex transport and sequestration into vacuoles. Two ABCC-type transporters (AtABCC1 and AtABCC2) are involved in the transport of metal-phytochelatin complex into vacuoles. Park et al. (2012) reported that over-expression of AtABCC1 in Arabidopsis resulted in enhanced Hg(II) accumulation and tolerance. Iglesia-Turiño et al. (2006) also reported Hg-induced synthesis of PCs in Brassica napus roots. Dago et al. (2014) concluded that total PCs contents were directly related to Hg level in soil. These findings highlighted the role of PCs in capturing and fixing Hg from soils.
4.8.3 Mercury Chelation by Glutathione
Plants have established a very effective mechanism to contest with toxic effects of heavy metal. Under metal stress conditions, plants can generate low molecular weight thiols that induce
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high binding affinity for heavy metals (Bricker et al. 2001). Glutathione (GSH), sulfur containing tri-peptide, is the most important low molecular weight biological thiol. Various studies showed changes in levels of reduced glutathione GSH and oxidized glutathione (GSSG) after plant exposure to different heavy metals (Schűtzendűbel and Polle 2002). Glutathione participates in scavenging H2O2 into plant cells (Flora 2011). The GSH-triggered detoxification of ROS starts at an early stage of plant development (Cairns et al. 2006). The detoxification of ROS by GSH is via ascorbate–glutathione cycle. Glutathione S-transferase catalyses the binding of GSH with Hg to form Hg-GSH complex and helps them to sequester into vacuoles (Yadav 2010).
The contents of GSH and GSH to GSSG ratio was reported to increase up to a concentration of 40 μM Hg and then greatly decreased at 50 μM Hg in cell cultures of Sesbania drummondii. In this study, significant increase in GSH level suggests the potential of Sesbania drummondii to tolerate Hg stress (Israr and Sahi 2006). This implies dynamic role of GSH in detoxification of ROS. This was in accordance with those observed with other plants like Thlaspi caerulescens (Freeman et al. 2004), Pteris vittata (Cao et al. 2004) and Brassica (Zhu et al. 1999) in which increased production of GSH enhanced Hg accumulation and tolerance. A significant oxidation of Hg-GSH was reported in alfalfa under Hg treatments (Ortega-villasante et al. 2007). An increase in GSH contents was positively associated with increasing level of Hg in Vallisneria spiralis and Bacopa monnieri (Sinha et al. 1996). Under metal stress conditions, the level of GSSG increases at the expense of GSH (Foyer and Noctor 2005). A high GSH to GSSG ratio is essential to endure the role of GSH as an antioxidant (Foyer and Noctor 2005).
4.8.4 Antioxidant Enzymes and Mercury Toxicity
Complex and effective antioxidant system, which detoxifies the ROS, is very important for plants in order to defend cellular organelles and membranes from ROS damage. Plants possess a well-developed enzymatic system to scavenge ROS and combat oxidative damage (Table 4.3). These enzymes include CAT, GPX, SOD, APX, GR etc. The antioxidant enzymes bind and deactivate the highly reactive free radicals of ROS. During ROS scavenging, antioxidants generally donate an electron to free radicals of ROS to form neutral and harmless end products such as water and oxygen. Plant defence mechanism needs energy for enzymatic detoxification or efflux of the Hg ions. In plants, ATPases and others chemiosmotic proton/ion exchangers act as efflux systems.
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Table 4.3: Mercury-induced activation of antioxidant enzymes in different plant species grown in
hydroponics
Enzymes Plant species Hg level Duration References
SOD, CAT, APX Triticum aestivumφ 0.32, 10.90, 30
mg kg-1 - Li et al. 2013
CAT, APX Cucumis sativus 0 to 500 µm 10, 15 days Cargnelutti et al. 2006
GR, SOD, APX Atriplex codonocarpa 0.05,0.1, 1 mg/L 28 days Lomonte et al. 2010
SOD, APX, GR Sesbania cell cultures 0–50 μM 21 days Israr and Sahi 2006
SOD, APX, CAT, POD Oryza sativa 12.5, 100 mM -- Chen et al. 2012
APX, SOD Zea mays 6, 30 µm 7 days Rellan-Alvarez et al. 2006
APX Medicago sativa 3, 10, 30 μM 1 day Ortega-Villasante et al. 2007
GR Oryza sativa 0. 2 mM 10 days Wang et al. 2009
SOD, POD, APX, GR Medicago sativa 10–40 μM 6, 12, 24, 48, 72 hr
Zhou et al. 2007
NADH oxidase, LOX, SOD, POD, CAT, APX, GR
Medicago sativa 0–40 μM 7 days Zhou et al. 2008
SOD, POD, CAT Lycopersicon esculentum
10-50 μM 20 days Cho and Park1999
SOD, POD, CAT, GPX Clitoria ternatea 1 µg/10 ml 20 days Priya et al. 2014
NADH oxidase, LOX, SOD, POD, CAT, APX, GR
Medicago sativa 10 µM 1 day Zhou et al. 2009
CAT, SOD Malting barley 0.1, 0. 2, 0.5, 1mM
7 days Song et al. 2012
SOD, CAT, POX, APX Triticum aestivum 2.5, 5, 10, 25 μM 8 h Sahu et al. 2012
SOD, POD, CAT Suaeda salsa 20 μg L−1 15 days Wu et al. 2012
SOD; superoxide dismutase, APX; ascorbate peroxidase, GR; glutathione reductase, POD; Peroxidase, LOX; lipoxygenase, GPX; guaiacol peroxidase, CAT; catalase, AsA; ascorbic acid, GST; Glutathione S-transferase, AAO; ascorbic acid. Φ Grown in soil.
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Superoxide dismutase is considered the first line of defence against ROS. The SOD reacts and converts superoxide radicals to O2 and H2O2. Greater activities of SOD under Hg stress have been reported in Bruguiera gumnorrhiza and Kandelia candel (Zhang et al. 2007), Sasbania drummondii (Israr and Sahi 2006), Horduem vulgare (Song et al. 2012) and Lupinus albus (Esteban et al. 2008). The POD also plays a significant role in plant defence against ROS (Zhou et al. 2007). The POD is reported to scavenge H2O2 generation, MDA elimination, resisting lipid peroxidation in cells and maintaining cell membrane redox status (Priya et al. 2014). The POX reduces H2O2 to H2O using various reductants available in cells (Foyer and Noctor 2005). Enhanced activities of POD have been shown in many plants upon Hg exposure (Wu et al. 2012; Priya et al. 2014). CAT can eradicate H2O2 and play a key role in elimination of O2
•− (Priya et al. 2014). Under Hg stress, greater activities of CAT have been reported in different plants (Lycopersicon esculentum, Clitoria ternatea, Medicago sativa, Malting barley, Triticum aestivum, Suaeda salsa) which present the role of CAT in Hg stress tolerance in these plants (Li et al. 2013; Cargnelutti et al. 2006; Song et al. 2012). The APX plays a little role in detoxification of H2O2 due to its sensitivity to Hg (Priya et al. 2014). The accumulation of H2O2 is restricted by APX, which converts it to H2O (Hernandez et al. 2010). An increase in APX activity has also been described in different plants (Atriplex codonocarpa, Sesbania cell cultures, Oryza sativa, Zea mays, Medicago sativa etc.) subjected to Hg (Lomonte et al. 2010; Ortega-villasante et al. 2007; Zhou et al. 2007; Alvarez et al. 2006). Glutathione reductase (GR) maintains GSH pool in plant cells, and is involved in the reduction of GSSG to GSH using NADPH. The Hg-induced activation in GR activity has been reported in Atriplex codonocarpa (Lomonte et al. 2010), Medicago sative (Zhou et al. 2008) and Oryza sativa (Wang et al. 2009).
4.8.5 Salicylic Acid and MercuryToxicity
Salicylic acid is stress related protein which acts as a signalling molecule. Majority of abiotic stresses which are involved in anti-oxidative responses in plants are interestingly SA-regulated (Khan et al. 2015). It suggests that the SA is an internal signalling molecule which interacts with ROS signalling pathway. For example, accumulation of H2O2 was reported in Arabidopsis, Nicotiana tabacum and Brassica nigra when they were treated with exogenous SA (Rao et al. 1997). The generation of H2O2 and the SA-mediated stress reactions are associated with each other in most of the cases (Drazic and Mihailovic 2005). Mercury was responsible for the generation of H2O2 which is proposed as an event of oxidative stress (Zhou et al. 2008). Surplus H2O2 may be toxic to plants (Gill and Tuteja 2010). However, an adequate level of H2O2 generation can be considered as signal activating various genes for oxidative or antioxidative responses to external environmental changes (Miller et al. 2008). Salicylic acid and H2O2 have a cross-talk and it is involved in the regulation of plant homeostasis to heavy metals including Hg (Freeman et al. 2005). In Medicago sativa, SA depressed the generation of H2O2 in Hg-treated roots by balancing H2O2 generator (NADH oxidase) and H2O2 scavengers (antioxidant defence systems, e.g. enzymatic and no-enzymatic components) (Zhou et al. 2009).
4.9 CONCLUSIONS AND PERSPECTIVES
This chapter focused on the environmental contamination of Hg in soil-plant systems and its entry into human food chain through plants. Mercury enters into soil-plants system through natural and anthropogenic sources. It enters in plant roots through specific, non-specific transporters and other channels (aquaporins) usually used for essential nutrients and water absorption. In plants, majority of Hg is accumulated in plant roots and sequestered in vacuoles through PCs and binding with cell wall. At lower levels, Hg inhibits seed germination, water absorption and various metabolic processes. It disturbs cell membranes and functionality of different enzymes. It also induces the production of ROS and causes damages to DNA and lipid peroxidation. Mercury volatilization can provide an ideal strategy to detoxify its immediate negative impacts on plants. For this purpose, plants can be engineered with higher expression of merB, that convert MeHg to Hg[II]. In addition, plants with ability to synthesize PCs that bind Hg[II] can further promote phytoextraction process. Majority of the wild plants species retains Hg
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in their roots, therefore, in future the genetic modification that improve Hg uptake into roots as well as its transport to shoots may provide a potential tool for phytoremediation of Hg contaminated areas and less accumulation of Hg in edible plant parts.
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CHAPTER-5
BIOGEOCHEMICAL BEHAVIOUR OF COBALT
IN SOIL-PLANT SYSTEM
5.1 INTRODUCTION
Cobalt (Co) is an imperative transition metal that lies between iron (Fe) and nickel (Ni) in the Periodic Table with two natural oxidation states (cobaltous Co2+and cobaltic Co3+). It possesses unique magnetic characteristic and has indispensable role as a cofactor in the production of vitamin B12 and enzymes depending on vitamin B12 for their functioning. It is rarely present in the Earth crust as compared with other trace elements (Kobayashi and Shimizu 1999). Adequate supply of Co2+ to ruminants (cow, sheep, goat etc.) produces resistance against parasites and infection.
Cobalt is not considered as an essential element for plants but promoting the development of plant growth (buds, plant stem, leaf discs and coleoptile etc.). Trace elements are important for the typical metabolic/physiological activities in plants and their availability at higher levels becomes toxic which damages various physiological and biochemical activities (Jayakumar et al. 2013).
Plants can absorb small quantity of Co2+ from soils and store in their different parts. Certain plant species extract metals from soils and help in phytoremediation. The phytoavailability and transportation of Co2+ in various parts of plants is species-dependent and regulated by variegated driving forces. Cobalt accumulation and storage in Triticum aestivum L. and Solanum lycopersicum L. were studied by Bakkaus et al. (2005). Active transport of Co2+ across cell membranes is involved for its absorption by plants. The Co2+ moves slowly from roots to shoots through xylem tissues on account of its low mobility (Palit et al. 1994; Barysas et al. 2002; Li et al. 2009). Cobalt sequestration in vegetal parts helps in unraveling its transport in higher plants (Bakkaus et al. 2005).
Like other heavy metals, Co poses a global threat to the environmental and human health. The exposure to Co2+ is associated with lung disease, especially in labor occupationally exposed to hard metal dusts (Sabbioni et al. 1994) and is also considered carcinogenic to humans (group 2B) by the International Agency for Research on Cancer. The anthropogenic activities like smelting, mining and soil amendments contribute to higher Co2+ concentration which may cause environmental risk (Cracan and Banerjee 2013). The possible unfavorable consequences of Co2+ to the community have needed to be examined.
The uptake of Co2+ at higher concentrations showed pronounced effects on the phytoavailability of essential nutrients, decreased net photosynthesis, destroyed integrity of chloroplasts and minimized activity of Fe-containing enzymes in leaves (Van Assche et al. 2001). The excess of Co2+ induces Fe deficiency and decreases the activity of CAT and chlorophyll which was direct reaction of reactive oxygen species (ROS) on enzymes (Luna et al. 1994; Seregin and Kozhevnikova 2006). In an investigation, Rauser (1978) and Rauser and Samarakoon (1980) revealed accumulation of carbohydrates in absolutely stretched leaves of Vicia faba seedlings exposed to Co2+ because of surplus photo-assimilate exports using vain loading.
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5.2 GLOBAL USES OF COBALT
Cobalt is considered important in trace amounts for sustaining life and is mainly present in oxides, sulfides, arsenides, and hydrates minerals. Cobalt is an integral part of water soluble vitamin B12 (also called cobalamin) which is involved in red blood cells formation, and typical performance of brain and nervous system. The typical ores of Co2+ are erythrite [Co3(AsO4)2.8H2O)] also known as red cobalt, cobaltite (CoAsS), linnealite (Co3S4) and chlorathite. Cobalt rich compounds are used in various industrial processes such as strong oxidative desulfurization catalyst (Co molybdate, Co sulfate) for petroleum resources, and for the synthesis of polyesters, detergents, plastic, paints, glass, hair dyes, ceramics, pottery, enamel, jewelry, Ni–Cd batteries, porcelain, cement, rubber, and chemical industries and pottery. Cobalt is unique ductile metal with additional malleable features and is produced as a byproduct in refinery and mining of copper (Cu) and iron (Fe). This property allows transformation of Co2+ into both thin wires and sheets. The Co2+ is involved in retaining physical and chemical properties of super alloys manufactured at high temperatures, e.g. aircraft engines (Harper et al. 2011). According to US-GS (2016), Co mine production increased by 46% from 675x105 to 1240x105 metric tons during 2006 to 2015 (Fig. 5.1).
5.3 COBALT LEVELS AND SOURCES IN SOILS
Cobalt does not occur naturally as base metal, but is a component of more than 70 minerals such as arsenide, sulfides, hydrates, sulfo-arsenides and oxides (Donald and Barceloux 1999). Cobalt concentration in soil is mainly acquired from parent materials. The average value of Co2+ in the Earth crust is 10 mg kg-1. Different values of natural background concentration of Co2+ have been reported in literature, i.e. 6.9 mg kg-1 (Kabata-Pendias and Pendias 2011), 7.1 mg kg-1 (Eriksson 2001), 9.1 mg kg-1 (Burt et al. 2003), 10.4 mg kg-1 (FOREGS 2005), 17 mg kg-1 (Licht 2005) and 18 mg kg-1 (Takeda et al. 2004). The mean value for Co2+ in Lithuanian soils was 3.4 mg Co2+ kg-1 soil, and in some Russian soils was 5.5 mg Co2+ kg-1 soil (Kabata-Pendias and Pendias 1999). Swedish cultivated soils contain 0.4–14 mg Co2+ kg-1 soil with mean value of 7.1 mg Co2+ kg-1 soil (Eriksson 2001). The Co2+ concentration in South African soils ranges from 1.51 to 68.5 mg kg-1 (Herselman et al. 2005).
Figure 5.1: Annual world refinery production of cobalt (USGS 2016).
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Relatively high Co2+ concentration (up to 520 mg kg-1) was reported in serpentine rocks and up to 85 mg kg-1 surrounding ore deposits. The Co2+ ranges from 5.5 to 29.9 mg kg-1 and 5.5–97 mg kg-1 in soils of the United States and China, respectively. Soils derived from mafic rocks (rich in Mg and Fe) and clay deposits contain the highest amounts of Co2+. In Australian Feral soils (122 mg kg-1) and Japanese soils (116 mg kg-1) is associated with high Co2+ concentration (Kabata-Pendias and Pendias 1999). Typically, the highest concentration of Co2+ is found in loamy soils and sometimes in Histosols. The Co2+ levels and its dissemination in soil profile differ due to variations in soil-forming processes in various agro-climatic zones. It is present in small amount in soils containing large particles derived from glacial deposits under temperate humid continental climatic conditions.
5.4 PHYTOAVAILABILITY OF COBALT IN SOILS
The distribution of Co2+ in soil profile depends on a number of factors. However, oxides and hydroxides of Fe and Mn show a great affinity for selective adsorption of Co2+. This was examined that distribution of Co2+ in soils had a common similarity between the levels of Fe and Co2+ in soil profiles. The Co2+ concentration in soil often remained higher in the B horizon where higher concentration of Fe was found. Nevertheless, in Mn dominated soils, the relationship of Co2+ with Mn controls other aspects governing the Co2+ distribution. The accumulation of Co2+
may reach up to 300 mg kg-1 in Fe nodules. The reactivity of different Fe and Mn adsorbents is a function of their surface area, not their abundance.
Typical oxidation states of Co are +2, +3 dominated by more stable +2 state which forms
stable complexes with other ligands and complexes of Co[III] (Kabata‐Pendias and Sadurski 2004). Cobalt is an important constituent of animal nutrition and its adequate supply to grazing animals, particularly in ruminant livestock has been a matter of great concern. In ruminant animals, Co2+ of diet is incorporated into vitamin B12 by rumen microbes. Exposure to Co alone has allergic reaction such as dermatitis and asthma. The production of vitamin B12 (hydroxocobalamin) by rumens drops off quickly due to deficiency of Co2+ in their diet, thereby affecting its efficiency and digestive system of ruminants. Pasture soils in the North-Western Coast of Egypt having AAAc-EDTA extractable Co2+ in the range of 3.60-4.69 mg kg-1 (Nasseem and Abdalla 2003). It was found that extractable Co2+ level (<2.45 mg kg-1) in pasture soils may yield a dearth of Co2+ in animals. The typical practices followed to overcome the Co2+ shortage in ruminants are through the application of sulfate of Co2+ or EDTA chelates onto soils.
5.5 SOIL PROPERTIES ADN COBALT PHYTOAVAILABILITY
The Co adsorption by soil constituents depends on many specific and non-specific adsorption mechanisms. Cobalt speciation in soils relies on a number of factors, among which the most important is the reduction potential (Eh). Soil strength is one of the basic challenges of modern science which implies the devaluation in the pollutant mobility in soil-soil solution-plant systems for ecological stability (Anisimov et al. 2015). There may be solubilization of precipitated or adsorbed Co2+ due to a decline in soil pH and Eh. Anisimov et al. (2015) suggested a methodological approach for estimating contribution of soil characteristics in inactivating metals in soil-plant system. The modification in oxidation state of Co2+ from +2 to +3 by Mn oxides is trivial in soils. Cobalt was adsorbed/absorbed by Mn oxides due to interchange reactions resulting into the production of hydroxyl species that precipitated at the oxide surface. A variety of redox reactions are suggested for the sorption of Co2+ by Mn oxides. Non-significant correlations were found between Co2+ adsorption onto soil particles and extractable Fe and Mn in soils from Tanzania and Denmark which might be due to low concentration of Fe and Mn in soils (Borggaard 1987). Cobalt adsorbed by Fe and Mn oxides was well predicted. In soils, Co2+ sorption by Fe and Mn minerals increased remarkably with pH and was also driven by soil redox potential. In sludge, there were hostile interactions among Co2+ and metals (Ni, Cu, Zn, Ca, Fe) (Osuna et al. 2004). It was revealed that montmorillonite and illite are important clays due to their larger surface area, higher sorption capacity and comparatively easy release of Co2+.
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The Co phytotoxicity and phytoavailability to Solanum lycopersicum, Brassica napus and Hordeum vulgare from ten soils varying in their properties were studied by Li et al. (2009). They found that exchangeable Ca2+ was the main consistent single predictor for Co2+ while solubility of Co2+ was the key factor for its mobility and toxicity to plants. Application of soil amendments (farm yard manure) and variation in soil chemical properties affect the phytoavailability of Co2+ (Robinson et al. 1999). The impact of soil OMon metal availability varies and relies on OM contents and pH of soils. The presence of organic ligands activates adsorption of Co2+ onto Fe-coated silica compounds (Brooks and Herman 1998). The mobility of Co2+ is highly related to soilOMOM, and organic chelating agents which help in its movement.
5.6 SOIL - PLANT TRANSFER OF COBALT
The accumulation of Co2+ in plants is determined not only by its contents in soils but also by soil factors (soil fertility, acidic-alkaline nature, oxidation-reduction conditions and organic matter contents) and plants ability to absorb Co2+. In higher plants, Co2+ absorption by roots occur through active transport. Nevertheless, Co is predominantly transported by transpiration induced movement in the xylem tissues. Phytoavailability of Co2+ was mainly determined by its complexes with organic compounds. Bakkaus et al. (2005) determined the pattern of translocation and accumulation of Co2+ with other metals (K, Ca, Fe) in Lycopersicon esculentum and Triticum aestivum species. In plants, a slow mobility of Co2+ from roots restricted its translocation from stems to leaves. Application of Co2+ chelates and sulphates in solid and liquid form in pastures showed increased Co concentration in plants due to its more phytoavailability with chelates. Its species soluble in AAAc-EDTA extractants are significantly correlated with its levels in plants (Sillanpää et al. 1992). Bhattacharyya et al. (2008) indicated that Oryza sativa absorbed relatively easily the Co2+ that was sorbed onto Mn and Fe oxides in cow dung manure and municipal solid waste compost. But Co contents in straw and grains of Oryza sativa were not correlated with its contents in carbonate materials, residual fraction of grains and straw or organic matter contents.
Literature shows that soil enrichment with Co2+ raises its concentration in plants. The effects of excess Co, Cr and Cu were investigated by growing Brassica oleracea in refined sand where visible effects of applied Co2+ appeared first and were more pronounced as compared with those of Cu and Cr (Chatterjee and Chatterjee 2000). Translocation of Cr from roots to shoot was the lowest while that of Co2+ was maximum, and each applied heavy metal inhibited the translocation of other macro and micronutrients. The Co2+ behaves like other trace metals during its absorption and is transported with organic compounds in complexes, with a –ve overall charge (Wiersma and Van Goor 1979). Foliar applications of Co2+ are effective in correcting Co2+ deficiency while Co2+ can easily be absorbed by leaves cuticles. Liming and fertilizer applications affect solubility of Co2+ resulting in decreased phytoavailability (Klessa et al. 1989). Positive effects of Co2+ on legumes growth for N-fixation have been reported extensively but fodder crops are affected due to low Co2+ contents condition which leads to Co2+ deficiency in ruminants.
5.6.1 Molecular Understanding of Cobalt Absorption by Plants
The distribution of Co2+ in plants completely depends on the nature of plant species, e.g. it is hardly detected in Raphanus sativus while its significant amount has been found in the leaves of Vicia faba. However, leafy plants contain comparatively greater Co2+ concentration, e.g. > 0.6 Co mg kg-1 in Lactuca sativa, Brassica oleracea and Spinacia oleracea. While in forage plants, there is a wide variation in Co2+ contents (0.6-3.5 mg kg-1). There is a significant +ve correlation between applied concentration of Co2+ and that absorbed by forage crops (Mitra 2015). Cobalt accumulated in root cells, but free Co2+ decreased hydrolysis of Mg-ATP and proton transport in Zea mays root tonoplast vesicles. Free Co2+ ions regulate the system and influence the structure of lipid phase in tonoplast even when ATP complexes of Co2+ inhibit proton pumping. This inhibition of proton pumping can be clarified by interactions with protogenic domain of membranes through indirect process.
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In a study, two types of higher plants which differ in their strategies to absorb Fe and grow in very different ways, were investigated by Bakkaus et al. (2005) to get insight of Co2+ distribution in plants. In Triticum aestivum leaves, Co2+ translocation followed almost the same pattern as Ca and K through parallel network veins. Whereas, distribution pattern of Co2+ varied from that K, Ca and Fe in Solanum lycopersicum leaves which might be due to natural defense mechanism in plants through which Co2+ is sequestered into vacuole differently from other elements, but needs to be verified.
5.6.2 Competition between Cobalt and Cations for Absorption
Plants showed Fe deficiency due to high levels of trace elements and it was reported at all the combinations of Co2+ and Zn concentrations in Vicia faba (Hunter and Vergnano 1953; Wallace and Abou-Zamzam 1989). Enormous amount of accumulated metals in shoots and particularly in roots, decrease Fe uptake and compartmentation in these tissues, led by necrosis, decrease in CAT activity, and increase in non-reducing sugar contents in Hordeum vulgare (Agarwala et al. 1977; Sharma et al. 1977). Additionally, Co2+ in excess, inclines to interact synergistically with Zn (Berry and Wallace 1981). The perceived synergism is occasionally associated with mineral shortage (Wallace and Abou-Zamzam 1989).
Moreover, poor interactions between soil pH and plant Co2+ absorption or total soil Co2+
contents, and a number of other soil chemical characteristics have been examined to influence Co2+ uptake (Adams and Honeysett 1964; Adams et al. 1969; Berrow et al. 1982; Hill et al. 1953; Li et al. 2004). A +ve correlation between Co2+ absorption and soil organic carbon by ryegrass was noted for an array of Scottish soils (at acidic to neutral pH) whereas such correlation could not be established in Trifolium pratense (McLaren et al. (1987). A significant +ve correlation of Co2+ translocation in plants was reported for clayey soil of Canada and New Zealand but there was weak correlation for Scottish soils. Likewise, there was -ve correlation between Fe concentration and Co2+ absorption by plants and was markedly different where fertilizers were applied (McLaren et al. 1987).
Phytoavailability of Co2+ was observed during the past half century because it was crucial for ruminants’ physiology and symbiotic N- fixation. Research carried out in New Zealand, Australia, and Scotland has revealed positive effects of Co2+ on leguminous crops (Klessa, et al. 1989). The role of Co2+ on legumes growth is currently well established but its significance regarding other plant species is comparatively ambiguous. It has not been categorically studied that higher plants need considerable quantity of Co2+ but there are indications that an increased supply of Co2+ increased plant growth and development (Collins and Kinsela 2011).
5.6.3 Cobalt Sequestration into Plant Roots
The phytoavailability and transportation of Co2+ into various plant parts is governed by different processes and is species-dependent. Plant roots absorb Co2+ by active transport which is energy consuming process and its distribution may involve formation of organic complexes. Movement of Co2+ from roots to shoots is restricted due to its low mobility (Bakkaus et al. 2005; Barysas et al. 2002; Palit et al. 1994).
Nitrogen fixation by root nodules of leguminous crops is promoted by Co2+ being important component of vitamin B12 and cobamide coenzyme. Smith and Carson (1977) reported information on effects of Co2+ from several sources for non-legumes, but the final outlook is questionable. Even though, little amounts of Co2+ have been noticed in coenzymes of non-legumes, where it is not identified whether these complexes might have formed from microorganisms linked with plants (Nicholas 1975). Favorable effects of low Co2+ concentrations on plant biological metabolism are not yet fully explained. Apparently, several properties are cross-linked with relations to other trace metals (Kabata-Pendias 2010)
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5.6.4 Cobalt Translocation to Shoots
Cobalt toxicity symptoms are white, deceased/dead tips and leaves margins because of excess Co2+ absorption by roots, followed by translocation of H2O in transpiration stream which produced high Co2+ concentrations at leaf tips and margins. However, the prime response of plants to an excess Co2+ concentration is interveinal chlorosis of younger leaves, closely associated with lime-induced chlorosis. Anatomical and physiological effects of excess Co2+, like prevent damage to chromosome and mitosis, disordered phloem of the minor veins and endoplasmic reticulum of root tips have also been reported (Kabata-Pendias 2010).
5.7 FOLIAR ABSORPTION OF COBALT
Foliar applications of Co2+ indicated a substantial growth in root length, dry matter contents of clippings, total dry matter, increase in nod factor, and the amount of N2 fixed (Vraný 1978). The diverse properties of foliar Co2+ use on yield components, particularly, number of flowers and inflorescence can be linked with plant growth variations and foliage (%) cover through Co2+ treatment. A significant production in stem number of Viola reichenbachiana during 2011 is partially ascribed to plant growth and development at the 1st cut with foliar Co2+ application compared to other species, Viola had a greater (%) of foliage cover (Tomic et al. 2014).
5.8 TOXIC EFFECTS OF COBALT ON PLANTS
In plants, Co2+ was not categorized as an essential element while it was generally defined as beneficial (Marschner 2011). Excess concentration of Co2+ in plant tissues causes permanent destruction to cell membranes and plant cells, which declined plant biomass (Fig. 5.2, Ann. Table 4), plant growth, water and nutrients absorption, and ultimately cell toxicity (Pandey et al. 2009; Li et al. 2009; Karuppanapandian and Kim 2013).
5.8.1 Cobalt Toxicity to Plant Growth
Tewari et al. (2002) observed Co2+ toxicity effects similar to those of Fe deficiency as well as excess of Co2+ concentration in plants. However, very rare data are available regarding the toxic effects of excess Co2+ in plants. Phytotoxic studies of Co2+ in Brassica napus L., HordeumVulgare L. and Lycopersicon esculentum L. have revealed its harmful effects on biomass yield and morphological growth (Li et al. 2009). Results indicated that higher concentration of Co2+ restricted the accumulation of Fe, total protein, chlorophyll content and CAT activity in Brassica oleracea L. leaves (Hajiboland and Amirazad 2010). Moreover, high levels of Co2+ also affected adversely the transportation of Mn, Cu, Zn, P and S from roots to leaf tips of Brassica oleracea. Compared to excess Cr or Cu, Co2+ markedly decreased water potential in leaf air spaces affecting adversely the transpiration rate of leaves. When prolonged exposure to high Co2+ concentration, Brassica oleracea leaves exhibited higher water contents and diffusive resistance (Chatterjee and Chatterjee 2000; Luo et al. 2010).
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Figure 5.2: Biogeochemical behaviour of cobalt in soil-plant system.
Jayakumar et al. (2013) reported improved chlorophyll contents under low applied level of Co2+ which was responsible in increasing the growth of different plant organs, but excessive amount of Co2+ resulting in interference with the synthesis of photosynthetic pigment contents. The development of chlorophyll depends mainly on Fe concentration (Brown 1961; Briat et al. 2007). Kropat et al. (2000) suggested protoporphyrin as a precursor for chlorophyll synthesis. However, elevated level of Co2+ restricted the integration of Fe in protoporphyrin which decreased chlorophyll contents in leaves. This was reinforced by the finding that amounts of heavy metals such as Cu (Mocquot et al. 1996; Drążkiewicz et al. 2004), Co (El-Sheekh et al. 2003; Morrissey et al. 2009), Ni (Vijayarengan and Dhanavel 2005; Chen et al. 2009) induced chlorosis in plants, which were typically comparable to Fe deficiency chlorosis. Heavy metals impaired photosynthetic pigments due to the interference to protein, while treatments apparently obstructed the activities and synthesis of enzymatic proteins leading to biosynthesis of chlorophyll.
Co2+ in soil
adsorbed by Mn, Fe oxides
Corrinoids, B12(adenosylcobalamin
[AdoCbl]) and cobalamin (Cbl)
Figure: Movement of Cobalt in Soil-plant System
Effected by soil pH, SOM and reduction
potential Eh
Cobalt induce ROS
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5.8.2 Cobalt Genotoxicity
Cobalt behaves as a micronutrient, but its high concentration causes genotoxicity. The Co2+ compounds produce tumors in animals and are carcinogenic to human health. The excess level of Co2+ can cause polycythemia, atherosclerosis, gastric disturbances, ischemia and hyperglycemia, and reproductive problems (Valko et al. 2007). Cobalt is a chemical element having greater potential of allergic dermatitis. In beer drinkers, the beer containing Co2+ for stabilizing its foam may cause cardiac disease and respiratory sensitization (Rai et al. 2001; Hussain et al. 2012). In mammalian cells, Co2+ Co compounds induced changes in DNA, exchanges in sister chromatid and breaks in DNA strands, aneuploidy and peripheral lymphocytes, but weekly mutagenic to chromosomal abnormalities (Beyersmann and Hartwig 1992; Kawanishi et al. 2002) or substantial increase of DNA relocation when single cell gel electrophoresis was used (De Boeck et al. 1998). An increased level of Co2+ break DNA strands in cervical cancer cells, a small number of 6-thioguanine-resistant mutant and a more prominent frequency of SCE in V79 Chinese hamster cells (Barysas et al. 2002).
The hazards of Co2+ toxicity could be more alarming when worked with other mutagens. So, Co2+ makes alterations in chromatin DNA which is extracted from human cultured cells in the presence of H2O2 (Nackerdien et al. 1991; Baldwin et al. 2004). Cobalt restricts healing of the DNA, impaired by direct genotoxic agents such as X-rays, UV and alkylating substances (Hartwig et al. 1994; Hartwig 1998) even at small non-cytotoxic concentrations (Hartwig 1995; Tan et al. 2013). It is a mechanism of co-mutagenic effect of Co2+ on mammalian cells (Barysas et al. 2002).
Cobalt is a trace element for plants. However, its deficiency indicators are still unknown. In contrast, numerous studies on physiological actions of increased Co have already been reported but its multiple effects on different processes and functional systems included biosynthesis of RNA and DNA, inhibition of alkaloid and plant hormones (ethylene), ageing, crop growth and yield, cytokinesis, mitochondria respiration and response to stress conditions. Particularly, studies focused on the Co actions on photosystems-I and II, photosynthetic activities and photosynthetic pigments (Palit et al. 1994; El-Sheekh et al. 2003). However, Co2+ with other metal ions (Zn2+, Ni2+, Cd2+) produced a red-brown discoloration along veins of unifoliate leaves. While, later in stems and petioles of Phaseolus vulgaris, initial loss of leaf orientation and leaf drop was also examined of that plant (Rauser 1978; Rauser and Samarakoon 1980). However, Co2+ induced severe chlorosis in hydroponics study of Phaseolus vulgaris (Wallace et al. 1977; Singh 2005). The metals (Zn2+, Ni2+, Cr3+) concentration, soil pH, temperature, and plant species determine the bioaccumulation of Co2+ (Palit, et al. 1994). So, the Co2+ concentration in nine medicinal plants was significantly different. In Purearia tuberose and Alpina galangal, concentration of Co2+ was statistically different, i.e. 0.16 and 7.08 mg kg-1 in dry biomass, respectively (Rai et al. 2001). Such plants can be cultivated for removal of Co-pollutants from soils, i.e. for phytomining. For Co2+ metal, such plant is Berkheyacodii (Robinson 1999). In plants, little information about mutagenic action of Co2+ ions were gathered. In spite of negative results on mammalian cells, chromosomal aberrations were observed when plants exposed to Co2+
(Beyersmann and Hartwig 1992). The data regarding oxidative stress and enzymatic activities may be described by Co2+ action introduced in the oxidative stress response of plants (Tewari et al. 2002; Barysas et al. 2002)
The toxic potential of CoNP is important to evaluate because these are used in nanomedicine with nanoparticles of Ni and Fe, unitedly. Up to now, CoNP toxicity presented oxidative stress ( Papis et al. 2009; Monteiller et al. 2007), genotoxic effects (Colognato et al. 2008), inflammatory responses (Petrarca et al. 2005), and DNA damage (Ponti et al. 2009) during in vitro studies. A synergistic effect of Co2+ and tungsten carbide (WC) was reported in activating ROS production and inducing fragmentation of DNA (Lison et al. 1995). An increased concentration of Co is toxic to living cells both in vivo and in vitro, potentially comprising oxidative stress and inhibition of DNA repair (De Boeck et al. 2003; Shi et al. 2004; Unfried et al. 2007) while the genetic toxicity of Co2+ alone is still debatable. It is generally described that whether NPs can cause cytotoxic effects and pose severe threat to human health.
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5.8.3 Cobalt-Induced Oxidative Stress and Lipid Peroxidation
Catalani et al. (2012) showed that Co2+ ions produced ROS in vitro and in vivo studies. In a Fenton type reaction, Co2+ catalyzed the formation of hydroxyl radicals (•OH) from H2O2 (Table 5.1) (Semenza 1999; Valko et al. 2006). In addition to stimulation of hexose monophosphate shunt and DNA damage to ROS, prolonged exposure to Co2+ could develop oxidative stress, decrease GSH level but increase GSSG level (Valko et al. 2006; Lewis et al. 1991; Hoet et al. 2002).
Cobalt induced cell death by DNA damage, apoptosis and necrosis (Zou et al. 2002; Zou et al. 2001). The Co2+ can cause activation of caspases (Zou et al. 2002), enhanced phosphorylation of mitogen-activated protein kinases (Zou et al. 2002), DNA fragmentation (Graham et al. 2004), transcriptional repression of p53 gene, and increased production of ROS (Olivieri et al. 2001; Zou et al. 2001; Begović et al. 2016). In rats, neurotoxic effects of Co2+ have revealed behavioral changes and depletion of neurotransmitters such as serotonin, dopamine and norepinephrine (Czarnota et al. 1998). The Co2+ decreased post synaptic mechanism triggered by neurotransmitters and limited the synaptic transmission from a presynaptic obstruction of Ca2+ channels (Gerber and Gähwiler 1991).
The interactions of Co2+ with mitochondria and the formation of ROS (Begovic et al. 2016) likely cause mitochondrial neuropathies. However, specific mutations (reshuffling of genes) in the mitochondrial DNA may cause mitochondrial dysfunction and could increase the impact of some environmental stresses. This response is genetically strengthened with a potential role played by polymorphic variation of mitochondrial genetic background.
Table 5.1: Cobalt-induced enhanced production of ROS in various plant species grown in
hydroponics
ROS Plant species Co level Duration References
H2O2, O2- Brassica juncea 100 µm CoCl2 2, 4, 6 days
Karuppanapandian and Kim 2013
O2- Lemna minor
0.01, 1 mM CoCl2
3 days Begovic et al. 2016
H2O2, O2- Zea mays 1-10 mM CoCl2 - Myton and Fry 1994
H2O2, O2-
Solanum tuberosum
25 µmol L-1 CoCl2
30 days Li et al. 2005
H2O2 Triticum aestivum 50 μM CoCl2 6 hrs Wu et al. 2014
H2O2 Phaseolus aureus 50, 100, 200, 300, 400 µM CoSO4
30 days Tewari et al. 2002
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide.
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5.9 COBALT DETOXIFICATION IN PLANTS
The IWGotEoCRtH group (2006) mentioned that Co2+ as a transition element possesses a number of chemical characteristics with Fe because of which it may catalyze the disintegration of H2O2 by a Fenton-like reaction. Mostly the ROS are produced in the presence of mixture of Co2+ and H2O2 (Moorhouse et al. 1985; Lison and Lauwerys 1993), and the exact formation of radicals is still a matter of discussion. These ROS were responsible for genotoxicity using a deoxyribose degradation assay with a significant production of •OH when Co2+ (d50, 4 μm;6 μg/mL) were incubated with H2O2 (Lison and Lauwerys 1993). However, this consequence was less prominent than that perceived with same concentration of Co2+. The cobalt activity was enhanced almost three times when linked with WC particles.
In a cell-free system, using 5,5-dimethyl-1-pyrroline N-oxide (DMPO) spin trapping and electron spin resonance, Lison et al. (1995) stated that Co2+-metal (d50, 4 μm; 1 mg/mL phosphate buffer) particles formed small quantity of ROS, presumed to be •OH. This action was observed in the absence of H2O2 and could not be replicated with Co2+, signifying that a Fenton type reaction was not included. The production of ROS by Co2+ particles was enhanced (Co: WC, 6:94) in the presence of WC particles. Resultantly, solid-solid interaction between particles was observed in which transfer of electrons from Co2+ particles reduced O2 at the surface of WC, those are subsequently solubilized and oxidized. The resulting Co2+ did not interfere during the production of ROS. Further, the dispersive interaction between WC and Co particles showed that the association between induced or permanent dipoles works like a new chemical entity, with physicochemical characteristics different from those of the each constituent individually (Zanetti and Fubini 1997). Free radical formation initiates from activated oxygen produced at the WC surface. In the presence of H2O2, the WC-Co2+ complex reveals a peroxidase-like activity (Fenoglio et al. 2000; Vallyathan 2002).
5.9.1 Vacuolar Compartmentalization of Cobalt
At sub-cellular level, Co2+ caused starch accumulation in plastids those caused the greatest detrimental effects by producing amyloplasts, amylo-chloroplasts and chloro-amyloplasts from chloroplasts. Similar findings were also described in Zea mays L. (L'Huillier et al. 1996; Hummel et al. 2010) and Vicia faba L. (Nyitrai et al. 2004). However, under Co2+ stress, accumulation of starch may be due to (1) photosynthate inhibition from source to sink tissues leading to accumulation of monosaccharide and disaccharide carbohydrates which might function as substrates for starch biosynthesis, (2) in higher plants, inhibition of plant growth than inhibition of photosynthesis leading to an excess amount of carbohydrates that is likely to be kept as starch. Often, Co2+ toxicity in plants is produced by an inhibition of Fe absorption which causes Fe deficiency (Chatterjee and Chatterjee 2003; Sree et al. 2015).
5.9.2 Cobalt Chelation by Phytochelatins
In cellular signaling, heavy metals stress induced the activation of glutathione (GSH) (Cobbett 2000) and phytochelatin synthase (PC-synthase) (Shahid 2010). In response to metal stress, a stimulation of GSH synthesis pathway is vital for synthesis of Cys-residual rich peptides, and PCs that normally bind and detoxify toxic heavy metals. The PCs are (g-Glu- Cys)n-Xaa polymers, having 2–11 g-Glu-Cys replicates which work as high affinity chelators (Zhu et al. 1999). The PCs are produced non-translationally from reduced GSH in a transpeptidation reaction catalyzed by the PC-synthase (g-glutamylcysteine dipeptidyl transpeptidase) enzyme (Zenk 1996; Yadav 2010). The Co2+ is not recognized to stimulate PC-synthase (Rauser 1995) or its induction did not differ significantly(Vatamaniuk et al. 2000). While, the citrate and Cys contents were increased in the non-accumulators (Silene cucubalus and Rauvolfia serpentina) cells, and in the hyperaccumulator (Crotolaria cobalticola) under prolonged exposure to Co2+. This was examined by size exclusion chromatography that free Cys was performed in Co2+
complexation in plant cells (Kleizaite et al. 2004).
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5.9.3 Cobalt Chelation by Glutathione
In-vitro ROS production by Co[II] from H2O2 and associated oxidative damage to DNA have been reported by Mao (1996). The development of •OH or singlet oxygen (1/2O2) described that the oxidative potential of Co[II] could be amended by many chelators such as 1,10-phenanthroline and anserine. Shi et al. (1993) and Valko et al. (2005) reported that undergoing redox cycling reaction, the ROS production from Co[II] ions and hydroperoxides showed that several chelating agents, such as endogenous compounds, decreased glutathione (GSH), facilitated the generation of reactive oxygen and nitrogen species. Sarkar (1994) postulated that proteins or oligopeptides signify other ligands those can modify the redox cycling potential of Co[II] ions. The occurrence of histones in nucleus permitted formation of metal mediated ROS in biologically relevant molecules such as DNA basis. The ability of Co[II] to substitute Zn[II] in the DNA-binding domains of nuclear proteins (transcription factor) might allow the in-situ generation of free radicals those may affect genetic regulatory/response elements and may elucidate the carcinogenic and mutagenic potential of these metals.
5.9.4 Antioxidant Enzymes and Cobalt Toxicity
Plants have developed antioxidative defense mechanism to detoxify ROS. However, failure to quenching harmful ROS or the subsequent production of an oxidative chain of reactions lead to damage to plant cells (Shahid et al. 2014a). The defense mechanism of plants involves both enzymatic and non-enzymatic antioxidants. The antioxidative enzymes defense system is the first line of defense against ROS, and includes a number of enzymes such as SOD, POX, CAT, DHAR,MDHAR, and GR (Table 5.2) (Pourrut et al. 2011; López-Moreno et al. 2016). These enzymes are substrate-specific as SOD dismutases superoxide ions to H2O2, which itself is very dangerous for plant cells and is, in turn, neutralized to H2O by CAT and POX (Manoharan et al. 2005). The role of Co2+ in the antioxidative enzymes in plants is contradictory, and both activation and inhibition are reported for numerous crop species like, Brassica oleracea L. (Gopal et al. 2003; Pandey et al. 2009), Lycopersicon esculentum L. (Gopal, et al. 2003), Brassica oleracea L. (Pandey, et al. 2009), Vigna radiata L. (Tewari, et al. 2002) and.
Cobalt toxicity and detoxification mechanisms are well understood; however very little information is available about Co2+ signal transduction pathways. Cobalt toxicity has increased the activity of POD (Bisht and Mehrotra 1989; Agarwala et al. 1977). Contrasting results of CAT activity by Co2+ toxicity are reported in literature. The CAT activity has been increased at 50 M, whereas, decreased at 100 M concentration of Co2+. In other reports, decrease in activity of CAT by Co2+ toxicity is reported by Agarwala et al. (1977) and Bisht and Mehrotra (1989) and increased by others (Tewari et al. 2002; Hasan et al. 2011). Karuppanapandian and Kim (2013) reported in Brassica juncea L. that Co2+ treatment caused a significant decrease in CAT activity, as evidenced by high levels of H2O2 in plant cells. Similar findings were described in other crops treated with Co2+ like Brassica oleracea L. (Pandey and Sharma 2002), Solanum lycopersicum L. (Gopal et al. 2003) and Spinacia oleracea L. (Pandey et al. 2009). It has been suggested that decline in the activity of CAT by Co2+ toxicity might be the result of deactivation of enzymatic antioxidants, resulting from either direct interaction with metal ions or due to the increased level of Fenton type generation of radicals (Cavalcanti et al. 2007).
The role of mitogen-activated protein kinases (MAPKs) (activation of 44 and 46 kDa MAPKs) in signals transduction as a result of Co2+ toxicity intervened by ROS cannot be ruled out. Several studies confirmed the activation of MAPKs through ROS generated by heavy metals toxicity in various plants (Kovtun et al. 2000; Jonak et al. 2004). A possibility of activation of MAPKs activation by Co2+ toxicity independent of ROS production cannot be ignored and needs further investigations. The overall importance of MAPK signaling against heavy metals stress is, therefore, unclear and the evidence of MAPK pathway involvement in heavy metal stress has rarely been reported in plants.
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Table 5.2: Cobalt-induced activation of antioxidant enzymes in different plant species grown in
hydroponic
Enzyme Plant species Co exposure level
Duration Reference
CAT Lycopersicon lycopersicum
62.5,125, 250, 500, 1000 CoFe2O4
15 Days López-Moreno et al. 2016
POX, CAT, MDHAR, DHAR, SOD
Brassica juncea 100 µm CoCl2 2, 4, 6 days Karuppanapandian and Kim 2013
SOD Vicia faba 2.5, 7.5 mM Co(NO3)2 41 days Kleizaite et al. 2004
CAT, SOD, GSH, GPX Pavlova viridis 0-200 µmol/L CoCl2·6H2O
24 hrs Mei et al. 2007
SOD, APOD, POD, CAT
Phaseolus aureus 50, 100, 200, 300, 400 µM CoSO4
45 days Tewari et al. 2002
CAT, POX, PPO Glycine maxφ 50, 100, 150, 200, 250 mg kg−1
15, 30, 45, 60, 75 days
Jayakumar et al. 2008
SOD Lemna minor 0.01, 1 mM CoCl2 3 days Begovic et al. 2016
CAT, SOD, POD Lycopersicon esculentum
100, 200, 300 µM CoCl2 4, 8, 12 h Hasan et al. 2011
POX, CAT, PPO Arachis
hypogaeaφ 50, 100, 150, 200, 250 mg kg−1CoCl2
30 days Jaleel et al. 2008
POX, CAT, PPO Sesamum indicumφ
50, 100, 150, 200, 250 mg kg−1CoCl2
25 Days Jayakumar et al. 2013
CAT, APX, SOD, POD Solanum
tuberosum 25 µmol L-1 CoCl2 30 days Li et al. 2005
SOD, APOD, POD, CAT
Phaseolus aureus 50, 100, 200, 300, 400 µM CoSO4
30 days Tewari et al. 2002
5.10 HORMETIC EFFECT OF COBALT TOXICITY
Similar to other heavy metals, Co could induce toxicity to plants at higher applied levels. At lower levels, Co has numeral beneficial effects, especially in leguminous plants (Pilon-Smits et al. 2009). In a latest study conducted by Gad (2006) using Pisum sativum, the application of Co (8 mg kg-1) onto soils improved crop growth, nutrient levels, nodule formation and weight, as well as seed quality and pod yield. Other positive effects of Co uptake comprise of leaf senescence retardation through inhibition of ethylene bio-synthesis, and development of drought resistance in seeds. In medicinal plants, Co also activated isoquinoline accumulation (an alkaloid) by up-regulation of the biosynthesis of aromatic amino acid precursors of alkaloids. This mechanism suggests that Co could indirectly bring biotic stress resistance, but this theory has not been clearly described until now. In Co2+ hyperaccumulators, higher levels of Co2+ in tissues may also address direct security from pathogens or herbivores as was presented for other hyperaccumulated elements. This elemental defense needs to be investigated. Meanwhile, Co is
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important for animals and Co fertilization of crops will have the supplementary positive effect on plant nutrition.
5.11 CONCLUSIONS AND PERSPECTIVES
Studies show that cobalt has a unique magnetic characteristic and indispensable role as a co-factor in the production of vitamin B12 and enzymes. Cobalt is considered important in trace amounts for sustaining life and is mainly present in oxide, sulfide and hydrate minerals. Distribution of Co2+ in soil profiles depend on a number of soil factors. However, oxides and hydroxides of Fe and Mn show a great harmony with adsorption of Co2+. Studies have shown that Co2+, along with Zn, Mn and Ni, is less soluble and could decrease micro-pore diffusion into various soil components. The Co speciation in soils relies on a number of factors, among which the most prominent one is the oxidation-reduction potential. Accumulation of Co2+ in plants is determined not only by its contents in soils but also by other soil factors (soil fertility, acidic-alkaline nature, oxidative-reductive conditions and organic matter contents) and plants ability to absorb Co2+. Among soil factors, the most effective strategy for Co removal, from contaminated soils, is to increase soil pH and/or Mn concentration. However, integration of Co[II] into Mn hydrous oxides and its oxidation to Co[III] can efficiently decreases the absorption of the applied Co to plants.
Accumulation of Co2+ in plants at a higher concentration might decrease phytoavailability of other nutrients and photosynthesis, destroy the integrity of chloroplasts and minimize the activity of CAT. Moreover, Co2+ toxicity interferes with various metabolic processes and disturbs cell membranes and functionality of different enzymes. It also induces the production of ROS and causes DNA damage and lipid peroxidation. Cobalt metal and its ionic forms could cause genotoxicity by oxidative damage of DNA carried through ROS. To combat this Co2+ stress, plants activate numerous defence mechanisms including antioxidants, Co2+ sequestration into vacuoles by forming complexes, binding Co2+ with GSH, PCs, and amino acids.
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CHAPTER-6
BIOGEOCHEMICAL BEHAVIOUR OF NICKEL IN SOIL-PLANT SYSTEM
6.1 INTRODUCTION
Nickel (Ni) belongs to transition series and a member of group VIII B of the Periodic Table along with cobalt, iron, platinum, palladium and five other elements. Nickel was first discovered in 1751, by a Swedish chemist Ronstadt. It has an atomic number of 28 and atomic weight of 58.71. It is naturally occurring element in Earth crust as a structural component of several minerals (Kabata-Pendias 2011). It comprises of almost 3% of the Earth and is ranked as 24th the most abundant element in the Earth’s crust (Alloway 1995), i.e. almost two-times the abundance of Cu. It occurs in igneous rocks abundantly, complexed with iron or in free metal form. By weight, Ni is ranked as the 5th most abundant element after oxygen, iron, silicon and magnesium. It contains five stable and nineteen unstable isotopes. It occurs in oxidation states of -1, +1, +2, +3 and +4 depending upon the environmental conditions, but the most dominant oxidation state is Ni[II] (Cempel and Nikel 2006).
Nickel is nutritionally essential micronutrient and plays many key functions for several species of plants, microorganisms and animals, (Fabiano et al. 2015). The toxicity or deficiency of Ni can occur respectively when too high or too little Ni is absorbed by living organisms. It may become toxic at 100 mg kg-1 FW in plants and at 420 kg ha-1 in soils (Tucker 2005). The role of Ni at cellular levels is known, but its deficiency state in plants and human body has not been yet explained in detail.
Generally, Ni in surface waters and soil is lower than 0.005 and 100 mg kg-1 respectively (Kabata-Pendias 2011). Nickel and its compounds are released into the atmosphere from natural sources such as volcanic eruptions, windblown dust and from anthropogenic activities (ATSDR 2015b; Cempel and Nikel 2006). According to an estimate, natural sources release about 8.5 million kg of Ni each year into the atmosphere. These natural sources include volcanoes, vegetation and windblown dust (Schmidt and Andren 1980).
This metal is used in different industrial processes, like steel and alloy industry, electroplating, Ni–Cd batteries, mining and smelting. Almost five times higher Ni is released from the anthropogenic sources compared to natural sources (Nriagu and Pacyna 1988). Major part of Ni released into the environment comes from Ni–Cd batteries, raw material used in electroplating and metallurgical industries and catalyst used in food and chemical industries (Cempel and Nikel 2006). During the last decade, Ni pollution became serious environmental concern because its level elevated up to 26,000 mg kg-1 in polluted soils (Alloway 1995) and 0.2 mg L-1 in surface waters (Zwolsman and Van Bokhoven 2007) and. These Ni levels are about 20-30 folds higher compared to their natural levels in unpolluted areas.
Nickel holds the special place among heavy metals as unlike others (Pb, Cd, Ag, Hg etc.), Ni is a structural component of various enzymes including ureases, glyoxalases, peptide deformylases and methylCoM reductase (Fabiano et al. 2015) and, therefore, low level of Ni is essential for some of plant species. High levels of Ni exposure are linked with several toxic effects including induction of leaf necrosis, wilting and chlorosis (Drzewiecka et al. 2012), disruption of photosynthesis (Seregin and Kozhevnikova 2006), antagonistic effect with plant nutrients such as Ca, Zn, Cu, Fe, Mg and Mn (Skukla and Gobal 2009) and reduction in plant growth and development (Khoshgoftarmanesh and Bahmanziari 2012) and initiation of oxidative damage (Pietrini et al. 2015). At cellular level, Ni strongly impedes metabolic reactions and
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generation of the reactive oxygen species (ROS) in plants and thus could cause oxidative stress. Ni-induced over-production of ROS can induce changes in activities of antioxidant enzymes such as CAT, GPX, SOD, APX and GR (Maheshwari et al. 2009; Kanwar et al. 2012; Parlak et al. 2016).
The food chain contamination and transfer of Ni from the environment to animals and man take place through soil-plant link. Due to toxic effects of Ni on human and animal health, it has been recognised as a human carcinogen by institutions around the world (WHO/IARC 2012). Therefore, biogeochemical behaviour of Ni in soil-plant systems is of the substantial environmental significance. In soils, the persistence of Ni and decrease in its mobility include the reduction-oxidation phenomena, sorption, desorption, precipitation, complexation, and dissolution. Although all these processes/reactions can take place concurrently, yet adsorption mechanism is determining factor towards geochemical behaviour of Ni in soils.
6.2 GLOBAL USES OF NICKEL
Nickel has numerous useful properties and plays an important role in the modern technology and infrastructure developments. Nickel is increasingly used in production of stainless steel (≈ 60 % of global primary Ni use), alloys and electroplating of rechargeable batteries (ATSDR 2015b; Mudd 2010). Owing to increased demand of Ni in different industrial processes, Ni mining increased from 1.4 to 2.4x106 tons during the last one decade (USGS 2016). In 2011, Chinese stainless steel industry was the major user of primary Ni (5.21x105 tons), increasing demand of Ni by 22 % a year (USGS2011).
Elemental Ni is a silvery-white and malleable metal that has high resistant to strong alkaline conditions (US-EPA 1986). Nickel plated alloys and objects are widely used to manufacture kitchen and tableware, bathroom fittings, food processing, consumer white goods, fasteners, motor vehicles, jet turbines, cables and wires, surgical implants and ship building (DEPA 2005). The increased industrial demand of Ni has enhanced its extraction, mining and use by 6 % in different industrial applications (Rauch and Pacyna 2009). It is estimated that ≈ 1/3rd of raw Ni is obtained by smelting of lateritic ores, which represents 60-70 % of the global reserves.
Widespread extraction and use has caused significant increase in Ni levels in the biogeochemical cycles and enhanced human exposure through environmental pollution. The processing, production and recycling of Ni products resulted in Ni pollution of soils, waters and atmosphere where Ni levels range between 25 to 170 ng m-3 (Lippmann et al. 2006). Various species of Ni include sulfides, silicates, oxides and soluble compounds such as metallic Ni and chlorides. Nickel occurs in soil either in deposits of laterite type ores where main ore minerals are limonite, nickeliferous and garnierite (Butt and Cluzel 2013) or in deposits of magmatic sulfide with principal ore mineral is the pentlandite (Hoatson et al. 2006). The US-GS (2016) demonstrated 38% increase in Ni mine production from 15.80x109 to 2.53x109 tons during 2006 to 2015 (Fig.ure 6.1).
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Figure 6.1: Annual world refinery production of nickel (US-GS 2016).
6.3 NICKEL LEVELS AND SOURCES IN SOILS
Most significant Ni ores are Ni-Fe sulfide, Ni-Mg silicate and garnierite. The lowest Ni concentrations are reported in sedimentary rocks (lime stones, clays, shales and sand stones). On the other hand, the highest Ni contents exist in basic igneous rocks. Nickel level in soil greatly depends on the Ni concentration of the parent rocks. However, Ni soil levels also represent its pollution from anthropogenic sources as well as the soil-forming processes (Kabata-Pendias 2011). Nickel occurs in natural soils at trace levels except in serpentinic and ultramafic soils. Nevertheless, Ni2+ level in soils is enhancing at high pace in specific areas as a result of human activities such as emission from smelters, mining works, burning of oil and coal, phosphate fertilizers, pesticides and sewage (ATSDR 2015b; Murtaza et al. 2012, 2015). Soil Ni levels in contaminated sites may reach 20-30-folds (0.2–26g kg−1) higher than natural background range found in soils (10–1000 mg kg−1) (Izosimova 2005).
Soil Ni contents vary greatly and has been predicted to range between 3-1000 worldwide with a mean value of 22 mg kg-1 (Kabata-Pendias 2011). According to Seregin and Kozhevnikova (2006), the range of Ni in soils varies between 2-750 mg kg−1. According to Nriagu (1980), Eriksson (2001), Takeda et al. (2004), Licht (2005), Burt et al. (2003) and FOREGS (2005), the mean soil Ni concentrations are respectively about 16, 13, 26, 25, 19, and 37 mg kg−1. Similarly, Berrow (1982) analysed 4265 soil samples and reported a mean soil Ni value of 33.7 mg kg−1. However, according to Duke (1980), mean Ni level in soil is 86 mg kg-1 based on Ni contents in the Earth crust. In late nineteens, Chen et al. (1999) reported Ni level in rural soils of various countries in Canada (150 mg kg-1), Australia (60 mg kg-1), United Kingdom (60 mg kg-1), Netherlands (210 mg kg-1), Germany (200 mg kg-1), France (50 mg kg-1), China (20 mg kg-1), Japan (100 mg kg-1), South Africa (15 mg kg-1) and United State of America (420 mg kg-1). The soil-forming geology processes significantly affect the concentration of Ni in soils (Kabata-Pendias 2011).
Nickel is considered a widely-distributed pollutant because of its high soil contents worldwide (Hussain et al. 2013; Shahid et al. 2014g). Natural sources of Ni emission into the environment include weathering of ultrabasic igneous rocks, meteoric dust, forest fires, oil dust and sea salt (Nriagu 1979). The man-induced release of Ni is dominated by fossil fuel burning,
50
60
70
80
90
100
110
120
130
10
5T
ho
usan
ds m
etr
ic t
on
s
Nickel
109
municipal waste incineration, Ni mining and other primary production operations and high temperature metallurgical operations (Hussain et al. 2013; Yusuf et al. 2011). Moreover, Ni is also released by tobacco smoke, vehicular exhaust, and in-house smoke from cooking fuels and home-heating. High Ni contents in surface soils have been linked with anthropogenic activities especially agriculture and industrial activities.
6.4 PHYTOAVAILABILITY OF NICKEL IN SOILS
Nickel compounds are generally categorised into soluble and insoluble fractions with respect to their mobility and solubility in soils. Goodman et al. (2009) classified Ni compounds into four categories: oxidic Ni, sulfidic Ni, soluble Ni and metallic Ni. Nickel is reported to bind strongly with soil constituents, but to a lesser extent compared to Cu, Pb and Zn (Buekers et al. 2007; Moreira et al. 2010). Several factors affect Ni adsorption-desorption in soils which include pH, bulk density, texture, type and amount of clay, OM, and certain oxides and hydroxides (Mellis et al. 2004; Sabir et al. 2008; Soares et al. 2011; Shahid et al. 2014g). Oxides of Fe and Mn as well as soil clay contents are considered the most important soil adsorbents of Ni. Under alkaline soil conditions, Ni adsorption can be irreversible (Rai and Zachara 1984), which decreases its mobility and phytoavailability in these soils. Indeed, Ni is selectively sequestered by finely Fe and Mn oxides. Nickel competes with cations for adsorption onto soil sites and the presence of Ca2+ and Mg2+ can decrease its adsorption onto soils due to competition for binding sites. In addition to clay and oxides of Fe and Mn, Ni adsorption in soils strongly depends on soil pH and Ni concentration (Giusti et al. 1993).
6.5 SOIL PROPERTIES AND NI PHYTOAVAILABILITY
6.5.1 Soil pH and Ni Phytoavailability
Soil pH is the most important variable, which affects Ni geochemical behaviour in soil-plant systems. The solubility and phytoavailability of Ni in soils decreases with an increase in pH (Staunton 2004). The environmental concerns of Ni are linked with decrease in soil pH because of enhanced Ni absorption by plants. Soil pH is considered to be more important factor in controlling Ni adsorption-desorption compared to Fe-Mn oxides, clays and soil OM (Tye et al. 2004).
Free Ni ions activity in the soil solution has been correlated with Ni phytoavailability. Therefore, Ni absorption by plants is mainly dependent on soil pH, OM contents, and oxides of Fe and Mn (Rooney et al. 2007). Soil pH plays an important role in partitioning of Ni in soils, and this partitioning of Ni is an important factor governing its toxicity to plants in soil-plant system. However, in calcareous soils, soluble Ni may also depend upon the dissolution of the surface precipitates either on surface of clay minerals or onto soil carbonates. Precipitation and adsorption of Ni differ greatly in acidic and alkaline soils. The precipitation of Ni hydroxides may describe strong Ni partitioning in soils with higher pH (Li et al. 2011). Nickel absorption by Lathyrus sativus increased with a decrease in soil pH up to 5.0 and with an increase in pH up to 8.0 (Panda et al. 2007).
Soil amendments could change metals mobility in soils mainly by modifying its pH. Thus, soil amendments those could increase pH could be applied to decrease the availability of metals. Like other heavy metals, Ni is less mobile and phytoavailable under alkaline conditions compared to acidic and neutral conditions. This is attributed to enhanced precipitation of Ni as insoluble hydroxides and carbonates with increasing soil pH (Kiekens 1984). Nickel concentration usually in soil solution is found to be negatively related to soil pH (Anderson and Christensen 1988).
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6.5.2 Soil Organic Matter and Ni Phytoavailability
Organic matter may affect positively or negatively the availability of a metal via complex formation. McBride (1989) suggested metal binding strength with organic matter in the decreasing order of Cu > Ni > Pb > Co > Ca > Zn > Mn > Mg. Organic matter in soils has been of specific attention in sorption of Ni and other heavy metals by soils because of its substantial effect on CEC and more importantly the tendency of metal cations to make stable complexes with the organic ligands (Mellis et al. 2004; Sabiret al. 2008; Shahid et al. 2014g). Phytoavailability of Ni in soils is greatly affected by type and amount of OM that reacts with Ni, thereby forming chelates and complexes of variable stability (Shahid et al. 2014g; Lockwood et al. 2015). The Ni-organic association can occur both in soil solution and at the solid surfaces of soil constituents (Silveira et al. 2003). Some studies reported beneficial effects of using organic amendments on soil ability to store Ni and other heavy metals (Shahid et al. 2014g).
For Ni retention by soil solids, OMis important thus decreasing phytoavailability and mobility. Increase in Ni solubility is associated with the dissolution of humic acid constituents of OM. Usman et al. (2004) reported that sewage sludge applied OMsignificantly increased the mobile fractions of Ni, Zn, Cd and Cu. It was concluded that quality and type of OMgreatly affected the mobility of metals in soils. Ashworth et al. (2004) reported relatively poor retention of Ni by the sludge solid phase. Lavado et al. (2005) reported that after compost application, concentration of Ni, in Zea mays plants was considerably higher. The untreated organic amendments (city and industrial raw effluents) could cause a great increase in DTPA extractable Ni than that with composted organic amendments.
Nickel forms complexes readily with organic substances as organic acids and the other dissolved OM, which enhances Ni dissolution and desorption in soils. According to water solubility and plant absorption, Ni in soils can be divided into available, potentially available and unavailable pools. Plants absorb Ni at any time from the available pool. Relatively less data is available on plant Ni absorption compared to other micronutrients. Similar to other metals, plant Ni2+ absorption may involve both active and passive processes (Kochian 1991).
6.6 SOIL-PLANT TRANSFER OF NICKEL
6.6.1 Molecular Understanding of Ni Absorption by Plants
Plant absorption of Ni metal is accomplished by the active transport and passive diffusion but its relative contribution depends upon Ni concentration in soils (Fig. 6.2) (Seregin and Kozhevnikowa 2006). Active transport of Ni is more important at lower concentrations (< 34 mmol L-1) and diffusion process increases at high Ni concentrations due to its toxic effect (Jean et al. 2008). Soluble Ni has a limited potential to enter cells via divalent metal transporter 1 (DMT1). Nickel is reported to enter plant roots through calcium channels (Denkhaus and Salnikow 2002). The presence of Zn2+ and Cu2+ inhibit Ni absorption because of their entry into plant roots via same the transport system (Kochian 1991). Nickel can also enter plant roots via Mg ion transport system. In fact, Mg and Ni have the same size ratio and charge (Oller et al. 1997). Nevertheless, some studies reported no inhibitory effect of Mg on Ni absorption by plants (Cataldo et al. 1978). Maestri et al. (2010) concluded that different kinds of proteins transporter in the metal transport and the homeostasis include (i) tonoplast transporters are involved in metal uptake (ii) plasma membrane transporters (iii) vacuole re-mobilization (iv) xylem loading transporters and (v) endomembrane transporters. Nickel absorption by plants also depends upon the concentration of Ni2+ (Cataldo et al. 1978), acidity of soil solution (Antoniadis et al. 2008), plant metabolism (Aschmann et al. 1987), presence of other metals (Podar et al. 2004) and OM composition (Jean et al. 2008).
Nickel is reported to enter plant roots by two types of high affinity transporters (Eitinger et al. 2000) which include ATP-binding cassette (ABC-type transporter) and transition-metal permeases (such as HoxN of Ralstonia eutropha). The best characterized Ni ABC-type transporter is reported in Escherichia coli. The ABC-type transport system of Ni consists of five
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proteins, NikA–E. These proteins are involved in ATP-mediated absorption and transport of Ni. The NikA is a soluble Ni-binding protein, NikB and NikC form transmembrane pores by which Ni can pass; while NikD and NikE provide energy to Ni2+-transport by hydrolysing ATP. Under high Ni contents, transcription of NikABCDE is blocked by NikR protein.
6.6.2 Competition among Ni and Cations for Absorption
Nickel is usually present in divalent form (Ni2+) in the environment and is most available form to plants. Divalent cations such as Ni, Fe and Cu possess similar physicochemical characteristics and hence compete with each other in physiological and biochemical processes in plants (Ghasemi et al. 2009). Nickel can enter plant roots via symplast pathway by AtIRT1 transporters (Nishida et al. 2012). Schaaf et al. (2004) reported that plant can absorb Ni as Ni–phytosiderophore (Ni-PS) complexes by the ZmYS1 transporter. It was reported that coexistence of protons and cations such as Ca, Mg, Na, K can affect Ni uptake, absorption and toxicity due to competition (Kalis et al. 2006). It is known that metals in soil/solution can induce synergistic and antagonistic impacts on other nutrient absorption, deficiency and toxicity to plants. Due to competition for same absorption sites, Ni affected absorption of Mn, Zn, Fe Ca, Mg and Cu (Chen et al. 2009). Nickel is reported to increase absorption of S and P in Lolium perenne (Yang et al. 1996).
Figure 6.2: Biogeochemical behaviour of nickel in soil-plant system.
Soil Plant
Transporters
ZmYS1
AtIRT
ATP
binding
cassette
Ni2+
Ni2+
Ni2+
Ni2+
Ni2+Ni2+
Ni2+
Ni2+
Ni2+
Ni2+
Nickel exists as adsorbed or complex on
organic cation, inorganic crystalline minerals,
water soluble, free-ion or chelated metal
complexes in soil
NRAMP
ZIP
transporters
Enter plant roots via AtIRT1 transporters as
Ni-PS complexes by ZmYS1 transporter
Ni species in soil
Ni chloride, sulfide and silicates nickeliferous
limonite
Ni2+
Absorption through
calcium channels
Chlorosis and necrosis, damages metabolism
transpiration and photosynthesis, ultrastructural
alterations, ROS overproduction, lipid peroxidation
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6.6.3 Nickel Translocation from Root to Shoots
Unlike hyperaccumulator plants, sensitive plant species retain most of the Ni absorbed from soils in root cells. Once Ni is loaded into xylem, flow of xylem sap can efficiently and rapidly transports Ni towards shoots (Centofanti et al. 2013). Metal transporters transfer metal ions from root symplast to xylem apoplast, and this transport is generally driven by transpiration pumps (Álvarez-Fernández et al. 2014). Complete information on molecular factors governing translocation of Ni is still lacking. However, numerous plant metal transporter families, organic acids and natural chelators have been identified to enhance Ni absorption and its translocation from roots to shoots. Each transporting protein catalyses the transport of several ions but with different affinity. Specific transporters in plants responsible for Ni hyperaccumulation have not yet been recognized (Rascio and Navari-Izzo 2011).
Nickel is transported from roots to shoots by different metal specific transporters. Assunção et al. (2008) stated that preferred Zn absorption over Ni by Ni and Zn hyperaccumulator plants exposed to equal levels of both cations strongly recommended that Ni entry into roots and transportation from roots to shoots may be mediated by Zn transport system. It is reported that NRAMP and ZIP families play key role in the transport of Ni and other metals (Guerinot 2000). Mizuno et al. (2005) reported that NRAMP and ZIP transporter genes (TjZnt1, TjZnt2 and TjNRAMP4) from Ni hyperaccumulator Thlaspi japonicum were cloned and their Ni transportation ability was demonstrated by the expression in yeast. Further, the NRAMP and ZIP transporters isolated from Thlaspi japonicum, appeared to protect cells from higher levels of Ni. Transporter genes such as TjZNT1& TjZNT2 contain long two His-rich domains (HRDs) and act as strong Ni binding regions.
6.7 TOXIC EFFECTS OF NI ON PLANTS
Nickel is an essential micronutrient for plants. It is well-established that plants need adequate levels of Ni (0.01 to 10 mg g-1 dry wt.) to complete their life cycle (Gratão et al. 2008). Nickel is structural constituent of nine metallo-enzymes. Nickel is linked with nitrogen metabolism due to the existence of two Ni atoms in the active centre of urease. Therefore, presence of Ni in growth medium stimulates hydrogenase as well as root nodule growth and development (Harasim and Filipek 2015). However, Ni may impede numerous physiological, morphological and biochemical processes (Ann. Table 5) in plant cells when its exposure level increases to supra-optimal levels of 420 kg ha-I in soils and 100 mg kg-1 in plants (Yilmaz and Parlak 2011).
Plant resistivity to Ni varies with the type of plant species (Freeman et al. 2004). Critical toxic level of Ni for sensitive crop species is >10 mg kg-1 dry mass (DM), >50 mg kg-1 for moderately tolerant species and >1000 mg kg-1 for tolerant plant species (Yusuf et al. 2011). In contrast, depending on cultivars and plant species, Kabata-Pendias (2011) demonstrated phytotoxic range of Ni from 40 to 246 mg kg-1 dry weight of plant tissues. According to Mishra and Kar (1974), toxic level of Ni varies widely from 25 to 50 mg kg-1 of soil. Some plants have been identified for their ability to tolerate high Ni level and are termed as Ni hyperaccumulators. Nickel hyperaccumulators include Brassica rapa, Brassica oleracea as well as the Leguminosae such as Vicia faba and Pisum sativum (Kabata-Pendias 2011). At these levels, Ni severely influences the growth and development of plants, thereby inducing phytotoxic effects at multiple levels and to various tissues (Hussain et al. 2013).
High Ni concentration inevitably binds with organic macromolecules and also denatures them. Nickel is transported to plant roots by oxygen atoms either as hydrated cation or metal complexes of organic acids (Salt et al. 2002). Additionally, excess concentration of Ni affects absorption of other nutrients by plant roots, damages metabolism of plants, inhibits transpiration and photosynthesis and causes ultrastructural alterations (Ahmad and Ashraf 2011). Exposure of plants to high levels of Ni causes chlorosis and necrosis due to disturbance of metabolism and Fe uptake (Seregin and Kozhevnikova 2006). Higher concentrations of Ni can also inhibit cell division and impede plant growth (Seregin and Kozhevnikova 2006). The very common symptoms of Ni toxicity in plants are wilting, necrosis and growth inhibition. Higher Ni uptake
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induces a decline in the water contents of monocot and dicot plant species (Bhalerao et al. 2015). At cellular level, Ni strongly impedes plants metabolic reactions and generates the ROS and cause oxidative stress. The Ni-induced over-production of ROS can induce changes in antioxidant enzymes activities such as CAT, GPX, SOD, APX and GR (Maheshwari et al. 2009).
6.7.1 Nickel Toxicity to Plant Growth
Nickel is reported to be toxic for most plant species and affects protease, amylase and ribonuclease enzyme activities because of which seed germination and growth of various crops is retarded (Ahmad and Ashraf 2011). It has been reported that Ni affects the mobilization and digestion of food reserves (carbohydrates and proteins) in germinating seeds and thereby decreases root length, plant height, fresh and dry weight, pigment content and increased malondialdehyde (MDA) contents and electrolyte leakage ((Ahmad and Ashraf 2011; Siddiqui et al. 2011). Nickel stress has been stated to affect plant yield and photosynthetic pigments in Vicia faba (Ahmad et al. 2007). Nickel and NaCl combination in germinating seeds of Brassica nigra caused a significant decrease in growth, pigments, leaf water potential and activities of anti-oxidative enzymes, lipid peroxidation, H2O2 contents, photosynthetic machinery by increasing electrolyte leakage and proline level.
Exposure to Ni at low levels (up to 20 mg L-1) stimulated germination index and rate and shortened the time for achieving 50% germination. On the other hand, higher concentration of Ni (40, 50, 60 mg L-1) induced a noteworthy inhibition in seed germination and interruption in achieving 50% germination. Moreover, radicle and plumule lengths as well as fresh and dry weights of all the Helianthus annuus cultivars were markedly decreased after Ni exposure (Ahmad et al. 2009). Similarly, exposure to 0.2 mM concentration of Ni inhibited evidently Triticum aestivum shoot growth (Gajewska et al. 2006). Similarly, after 7-10 days of exposure to 0.43 mM Ni, roots of Nicotiana tabacum turned dark brown and plants growth was repressed (Boominathan and Doran 2002). Seed germination of Cajanus cajan decreased in 1.5 mM solution of Ni (Rao and Sresty 2000).
Other studies also reported that Ni accumulation in plants seriously affects plants yield by decreasing number of seeds per pod considerably (Tripathy et al. 1981). Plant growth and total plant biomass also decreased when plants were exposed to high levels of Ni (Shahid et al. 2014g) possibly due to decrease in leaf density and leaf blade area, thereby causing a decline in number of flowers and fruits (Seregin and Kozhevnikova 2006). Generally, Ni-induced decrease in plant yield and biomass could due to decrease in nutrients supply to reproductive parts (Ahmad et al. 2007).
6.7.2 Nickel Genotoxicity
Nickel is capable to induce genotoxic effects in plants. In plants, Ni produces some genetic abnormalities like DNA-protein cross links, DNA strand break, nucleotide excision, single gene mutations, micronuclei, nucleic acid concentration modification and cell transformations (Das and Dasgupta 2000). However, there exist little data regarding Ni-induced genotoxicity in plants. The Ni-induced enhanced production of ROS has been reported as a key mediator of Ni-induced genotoxicity. A modification of gene mutation and expression due to enhanced production of ROS and other cellular oxidants are critical in Ni-induced genotoxicity. Nickel-induced production of hydroxyl radical can react directly with all the constituents of DNA (Cameron et al. 2011). The H2O2-mediated fragmentation of DNA is an indicator of the ROS-induced genotoxicity. Another example of ROS-induced genotoxicity is the breakage of double-stranded DNA. Fargašová (2012) reported significant increase of chromosomal aberrations by Ni. Li et al. (2015) reported that Ni can induce heritable genomic methylation variations in Arabidopsis root tip cells. They reported that Ni caused an increase in cytosine methylation in 18S rDNA in the nucleolus.
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6.7.3 Nickel-Induced Oxidative Stress
Nickel toxicity can induce enhanced production of ROS resulting in oxidative stress (Dourado et al. 2015). The ROS such as H2O2, ½O2, O2
• −, HO•, peroxyl (RO2•) and alkoxyl (RO•)
radicals are natural by-products of different biochemical and metabolic processes of plant cells (mitochondria and chloroplast), however, under Ni and other heavy metals stress their production increases significantly (Shahid et al. 2014a). Nickel is not a redox active metal; therefore, it is unable to generate ROS directly. Several studies reported Ni-induced increase in antioxidant activities due to over-production of ROS (Table 6.1) (González et al. 2015). Enhanced production of ROS by Ni is facilitated by its ability to bind with peptides, amino acids and proteins. Nickel-induced enhanced production of ROS can induce lipid peroxidation, DNA damage, oxidation of proteins and damage to pigment contents.
Lipid peroxidation is considered an important cellular oxidative impairment under metal-induced oxidative stress (Mishra et al. 2006). It is demonstrated that exposure to Ni induced lipid peroxidation in several plant species, suggesting oxidative damage under Ni stress (Seregin and Kozhevnikova 2006). Increase in lipid peroxidation has also been found in Zea mays (Baccouch et al. 1998), black gram (Dubey and Pandey 2011) and Lemna gibba (Yilmaz and Parlak 2011). It shows (Table 6.1) that H2O2 and O2
• − are the most commonly studied ROS in literature. Plant exposure to high Ni stress can enhance the level of H2O2, ½O2, O2
• − and HO• (Gajewska and Sklodowska 2007). Application of Ni to plants also enhanced plasma membrane NADPH oxidase, which is involved in Ni-induced production of ROS. Nickel-induced increase in ROS contents has been found in Alyssum bertolonii, Triticum aestivum (Parlak et al. 2016), Ceratophyllum demersum (Andresen et al. 2013), Zea mays (Ghasemi et al. 2013) and Coriandrum sativum (Tagharobiyan and Poozesh 2015).
6.8 NICKEL DETOXIFICATION IN PLANTS
Over-production of ROS seems a common response of plants to different stress conditions including Ni toxicity at high exposure levels (González et al. 2015; Soares et al. 2016; Parlak et al. 2016). In order to maintain normal biochemical and metabolic processes and functions under stress conditions, a balance between ROS generation and degradation is highly vital. In plant cells, threshold level of ROS is maintained by a complex defence system consisting of enzymatic and non-enzymatic antioxidants (Sabir et al. 2015; Shahid et al. 2014a). However, which particular plant defence system will operate under Ni stress conditions depends on numerous factors including level of Ni, duration of exposure, type of plant species etc. These defence mechanisms identify and mediate stress signals to regulate their metabolic processes under stress conditions. Different approaches adopted by plants under Ni stress include chelation by GSH and PCs, compartmentation into vacuoles and Ni scavenging by antioxidative enzymes.
6.8.1 Vacuolar Compartmentalization of Ni
Vacuolar sequestration plays an important part in the plant metal tolerance and homeostasis (Shahid et al. 2014a). Plant vacuoles are regarded as the main sink for metals in plants and it has been revealed that PC-metal complexes are pumped into vacuoles (Flora 2011). Transportation and sequestration of metal-PC complex inside vacuoles include active transport mechanisms functioning in tonoplast. Heavy metals transport across the tonoplast is mediated by ATP hydrolysis directly by primary active transporters like P1B-ATPases and ABC type proteins or by secondary active antiporters such as transmembrane pH gradient (Krämer et al. 2007).
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Table 6.1: Nickel-induced enhanced production of ROS in various plant species grown in
hydroponic
ROS Plant species Ni level Duration Reference
H2O2, O2- Ceratophyllum demersum 0.01 µM 12 days Andresen et al. 2013
OH-, H2O2, O2- Zea mays 0.05 µM 2 days Kumar et al. 2007
H2O2, O2- Solanum nigrum 100 µM 28 days Soares et al. 2016
H2O2, O2- Triticum durum 100 µM 9 days Gajewska and Skłodowska 2007
H2O2, Brassica napus 0.5mM 10 days Kazemi et al. 2010
H2O2 Oryza sativa 20, 40, 60 µM 1 day Lin and Kao2005
H2O2 Raphanus sativusφ 10, 50, 100, 200mg kg−1
40 days Singh and Prasad 2015
H2O2, O2- Triticum durum 1 mM 3 days Hao et al. 2006
H2O2 Zea maysφ 100 mg kg−1 28 days Wang et al. 2009
H2O2, O2- Alyssum murale 200,700 μm 1 hrs Agrawal et al. 2013
H2O2 Zea mays 100 µM 14 days Kumar et al. 2007
H2O2, O2- Psidium guajava 300, 1000, 3000 µM 7 days Bazihizina et al. 2015
OH-, H2O2, O2- Luffa Cylindrical 0.1, 0.2, 0.4 mM 11 days Awasthi and Sinha 2013
H2O2, O2- Hyoscyamus niger 50, 100 µM 1 day Nasibi et al. 2013
H2O2 Hordeum vulgare 50, 100, 200, 400,800, 1000 µM
12 days Kumar et al. 2012
H2O2, O2- Luffa cylindrica
50, 100, 200, 400, 800 µM
7 days Wang et al. 2010
H2O2, O2- Vitis vinifera 1–5 mmol L–1 2 days Pavlovkin et al. 2016
H2O2, O2- Pennisetum typhoideum
50, 100, 150, 250, 350 mg kg-1
7 days Rughani et al. 2010
H2O2 Coriandrum sativum 100, 200, 500 μM 14 days Tagharobiyan and Poozesh 2015
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H2O2, O2- Hordeum vulgare 0.1, 0.5, 1mM 7 days Jócsák et al. 2008
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide. Φ grown in soil.
It is considered that the vacuolar Ni storage in leaves is the main biochemical and metabolic detoxification mechanism by comparing Ni compartmentation between the hyper-accumulating and non-accumulating plant species (Krämer et al. 2000). Efflux tonoplast transporters are involved in the transport of excess Ni metal present in cytosol into vacuoles against the electrochemical gradient (Kabata and Janicka-Russak 2011). Krämer et al. (2000) observed in their study that by sequestering Ni into their vacuoles, Ni accumulation and tolerance of Ni-hyperaccumulator Thlaspi goesingenseis was improved considerably. In Thlaspi goesingense plant, increased production of MTP1 tonoplast protein has been associated with enhanced tolerance to Ni, Co and Cd (Persans et al. 2001).
6.8.2 Nickel Chelation by Phytochelatins
In Ni tolerant plant species, homeostasis is attained by removing them from metabolically active cellular parts. In plants, this is achieved by complexing toxic metals with ligands and transporting them to vacuoles. Some natural plant ligands such as PCs are considered as the most important metal-binding ligands. Phytochelatins are known to bind toxic metals including Ni and then Ni-PC complex is transported to vacuoles in plant. Phytochelatins are cysteine-rich binding proteins which bind metals by forming mercaptide bonds (Flora 2011) and thereby detoxify metals in plants (Verbruggen et al. 2009). Plants synthesise PCs when exposed to high levels of Ni. Phytochelatins are synthesized in plants from GSH by phytochelatin synthase (PC-synthase) activity. Nickel-induced enhanced production of PCs is reported in several plant species such as Thlaspi japonicum (Mizuno et al. 2003) and Nicotiana tobacum (Nakazawa et al. 2001).
6.8.3 Nickel Chelation by Glutathione
Glutathione, a sulphur containing tripeptide, is the most significant low molecular weight thiol in plant and animal kingdoms having a formula γ-glutamatecysteine-glycine. Glutathione is found in all the living organisms, including the eukaryotic and prokaryotic cells. Glutathione level in plants varies from 0.1 to 10 mM (Meister and Anderson 1983). Glutathione acts as a redox buffer against metal-induced enhanced production of ROS. In this way, GSH plays a key role to defend different cell parts and organelles against biotic and abiotic stresses. In plants, GSH is found in two forms as oxidized glutathione (GSSG) and reduced glutathione (GSH). Glutathione is synthesized in plants by ATP-mediated reaction in which γ-glutamylcysteine (γ-EC) synthesis is catalysed by γ-glutamylcysteine synthetase by forming peptide bond between amino group of cysteine and glutamate carboxyl group.
In plants, GSH is involved in numerous essential metabolic functions such as detoxification of xenobiotics (Israr et al. 2011), Ni sequestration into vacuoles (Freeman et al. 2004) and defence against the Ni-induced ROS (Foyer and Noctor 2005). Glutathione also helps plants to regulate different metabolic and developmental processes such as flowering and cell division (Yadav 2010). Wang et al. (2009) suggested that Ni stress in Zea mays enhanced GSH poll and GSH: GSSG ratio. Freeman et al. (2005) proved that in Arabidopsis enhanced GSH contents were achieved by activation of GR, a key enzyme involved in the synthesis of GSH. Israr et al. (2011) reported Ni-induced production of GSH in Sesbania drummondii.
6.8.4 Nickel Toxicity and Salicylic Acid
Salicylic acid (SA) is a significant and powerful signalling molecule in plants. Salicylic acid provokes plant tolerance against several abiotic and biotic stresses (Horvath et al. 2007).
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Exogenous application of SA enhanced Ni translocation to aerial parts in Matricaria chamomilla (Kovacik et al. 2009). Kazemi et al. (2010) reported that SA enhanced activities of GPX, CAT and APX whereas decreased lipoxygenase (LOX), MDA and H2O2 contents under Ni stress in Brassica napus. This showed that SA is involved in the regulation of cell membranes and activation of antioxidants to alleviate oxidative damage caused by Ni-induced over generation of H2O2 and MDA. Salicylic acid as a vital molecule is reported to enhance the resistance of plants to metal stress, salt anxiety, UV-B radiation, water stress, ozone stress and temperature stress (Khan et al. 2015). Exposure to Ni enhanced GSH accumulation in four types of Thlaspi hyperaccumulators (Freeman et al. 2005), which is basically a defensive approach adopted by hyperaccumulators to defend against Ni-induced oxidative damage (Freeman et al. 2004). Similar results have been demonstrated for non-accumulator Arabidopsis under Ni stress (Freeman et al. 2005). However, there exists rare data regarding the effect of SA on reactions of the entire antioxidant defence system under Ni stress and need to be explored. Moreover, it is known that plant chloroplast is an essential cell compartment and the most important intracellular generators of ROS. Therefore, it is necessary to explore how exogenous application of SA regulates antioxidant defence system in chloroplast under Ni stress.
6.8.5 Antioxidant Enzymes and Ni Toxicity
Plants possess a number of antioxidant enzymes and molecules which protect them against metal-induced oxidative damages. Many studies reported Ni-induced oxidative stress and activation of antioxidant enzymes in plants (Table 6.2) (Hussain et al. 2013). Previous investigations demonstrated that effects of Ni on the antioxidant enzyme system have already been studied in different plant species (Tan et al. 2008). Nickel-induced increase in antioxidant enzyme activities has been reported for Triticum aestivum (Parlak et al. 2016), Zea mays (Ghasemi et al. 2013) and Sesbania drummondii (Israr et al. 2011).
Nickel exposure to plants at low levels (0.05 mM) and for a short duration resulted in enhanced activities of GOPX, GR, POD and SOD to scavenge ROS and protect plants from oxidative stress (Freeman et al. 2004; Gajewska and Sklodowska 2007). Nickel-induced over-production of ROS and activation of antioxidant enzymes may differ with exposure duration and type of plant species. Gajewska and Sklodowska (2007) concluded that Ni application at 100 µM for 3, 6, 9 days decreased the activities of SOD and CAT, but increased the activities of GOPX, APX and GSH-Px. However, after 14 days, Ni treatment (10, 100, 200 µM) decreased SOD and APX activities. It is reported that Ni-induced activation of enzymes varies with exposure duration and applied level of Ni. Similarly, Ni (0.5 mM) decreased CAT and POD activities in leaves of Brassica oleracea after eight days (Pandey and Sharma 2002). Papadopoulos et al. (2007) also reported similar results for SOD, POD and CAT in leaves of Hydrocharis dubia after three days of Ni exposure (0.5, 1, 2, 3, 4 mM). Several studies reported Ni-induced activation of antioxidant enzymes in Zea mays, Vigna Mungo, Jatropha curcas, Brassica napus, Raphanus sativus and Triticum aestivum (Dubey and Pandey 2011; Gajewska and Skłodowska 2007; Yan et al. 2008; Kazemi et al. 2010; Ghasemi et al. 2013; Siddiqui et al. 2013).
Superoxide dismutase is the first enzyme to rapidly convert O2.- radicals to H2O2
(Gajewska and Skłodowska 2007). The Ni-induced increase in the SOD activity was observed in Triticum aestivum (Parlak et al. 2016), Zea mays by (Kumar et al. 2007) and Eichhornia crassipes by (González et al. 2015). However, Boominathan and Doran (2002) reported decreased SOD activity in roots of Ni hyperaccumulator Alyssum bertolonii after Ni exposure. The CAT transforms H2O2 to O2 and H2O. Yan et al. (2008) observed that Ni treatment resulted in significant increase in CAT activities of plants while results of Rao and Sresty (2000) were that the CAT activity decreased in Cajanus cajan seedlings grown at higher concentration of Ni. Guaiacol peroxidase (GPX) is abundant enzyme and its constitutive activity in plant roots is much higher than that in their leaves. Ghasemi et al. (2013) reported that Ni induced GPX activities in leaves of Zea mays seedlings. Peroxidases (PODs) are commonly found in plants as several isoenzymes due to its multiple functions. It is reported that POD activity, a H2O2- scavenging enzyme, increased after Ni exposure to plants. Kumar et al. (2007) reported increased POD activity in Zea mays after Ni exposure. Gajewska and Sklodowska (2007) also reported that excessive concentration of Ni
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may induce POD activities in order to increase activation of other antioxidant enzymes and hence caused the scavenging of ROS. Ascorbate peroxidase (APX) is a key enzyme of the glutathione-ascorbate pathway. It eliminates peroxide by conversion of ascorbic acid into dehydro-ascorbate (Foyer and Noctor 2005). The highest concentration of APX activity detected in roots and leaves of Zea mays suggests that APX might play an important role in H2O2 removal from cells in roots and leaves of Ni stressed Zea mays (Ghasemi et al. 2013). The GR is an important enzyme to protect plants against the oxidative damage by maintaining a higher ratio of GSH:GSSG (Shahid et al. 2014a). Kanwar et al. (2012) reported increased GR activity in Brassica juncea after Ni exposure.
Table 6.2: Nickel-induced activation of antioxidant enzymes in different plant species grown in
hydroponic
Enzyme Plant species Ni exposure level Duration (days)
Reference
CAT Lycopersicon lycopersicum
62.5,125, 250, 500, 1000 µm 15 López-Moreno et al. 2016
POX, CAT, SOD, MDHAR, DHAR
Brassica juncea 100 µm 2, 4, 6 Karuppanapandian and Kim 2013
SOD Vicia faba 2.5, 7.5 mM 41 Kleizaite et al. 2004
CAT, SOD, GSH, GPX
Pavlova viridis 0 to 200 µmol/L 1 Mei et al. 2007
SOD, POD, CAT Phaseolus aureus 50, 100, 200, 300, 400µM 45 Tewari et al. 2002
CAT, POX Glycine maxφ 50, 150, 200,250 mg kg−1 15, 30, 45, 60,75
Jayakumar et al. 2008
SOD Lemna minor 0.01, 1 mM 3 Begovic et al. 2016
CAT, SOD, POD Lycopersicon esculentum
0, 100, 200,300 µM 0, 4, 8, 12 hrs
Hasan et al. 2011
POX, CAT, PPO Arachis hypogaeaφ 50, 150, 200,250 mg kg−1 30 Jaleel et al. 2008
POX, CAT, PPO Sesamum indicumφ 50, 150, 200, 250 mg kg−1 25 Jayakumar et al. 2013
SOD; superoxide dismutase, APX; ascorbate peroxidase, GR; glutathione reductase, POD; Peroxidase, LOX; lipoxygenase, GPX; guaiacol peroxidase, CAT; catalase, AsA; ascorbic acid, GST; Glutathione S-transferase, AAO; ascorbic acid. Φ grown in soils.
6.9 HORMETIC EFFECT OF NITOXICITY
The hermetic concept that low levels of a substance induce stimulation in growth while high levels induce toxicity is known for some heavy metals. Among heavy metals, Ni is unique because of its essential role in plant metabolism. In fact, Ni is a structural constituent of numerous enzymes including ureases (Chen et al. 2009) and therefore exposure to plants at low Ni levels induces positive effects. However, at higher applied levels, Ni can induce toxicity to plants including disruption of photosynthesis, induction of leaf necrosis, chlorosis and wilting,
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antagonistic effect with other plant nutrients and inhibition of plant growth and low yields. It was reported that low level Ni application (up to 20 mg L-1) significantly promoted germination index and rate, and increased radicle and plumule lengths. However, at a higher concentration (40, 50, 60 mg/l), Ni induced considerable decrease in seed germination and length of radicle and plumule (Ahmad et al. 2009).
Exposure to low levels of Ni generally induces beneficial biological impacts on plant growth and development. Plants need sufficient levels of Ni (0.01 to 10 mg g-1 dry wt.) to complete their life cycle (Gratão et al. 2008). However, when exposed to high Ni level (420 kg ha-I in soil and 100 mg kg-1 in plants), it may impede numerous physiological, morphological and biochemical processes in plants (Yilmaz and Parlak 2011).
6.10 CONCLUSIONS AND PERSPECTIVES
Nickel is a known persistent toxic pollutant of concern. Nickel has many useful physicochemical properties thereby high industrial uses. Nickel has low mobility and phytoavailability in soils. Soil factors (pH andOM) greatly influence Ni mobility and soil-plant transfer. The Ni plays an essential role in plant metabolism, but becomes toxic at high applied levels. Nickel absorption is mainly via roots. There is no Ni specific plant transporter and absorption mainly takes place via Ca and Mg channels. Like other metals, Ni also accumulates primarily in plant roots. At high levels, Ni impairs various morphological, physiological and biochemical functions at cellular level. Nickel is capable of causing oxidative stress via over-production of ROS. Under Ni stress, plants activate numerous defence strategies to avoid its toxicity. These defensive approaches include Ni sequestration into vacuoles by forming complexes, binding of Ni with GSH, PCs and amino acids. Moreover, numerous antioxidants are involved to combat increased production of the Ni-mediated ROS.
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CHAPTER-7
BIOGEOCHEMICAL BEHAVIOUR OF ZINC IN SOIL-PLANT SYSTEM
7.1 INTRODUCTION
Zinc (Zn) is a transition metal that occurs naturally in the Earth crust, and is reported as the 24th the most abundant element. Its background concentration ranges from 10–100 mg Zn kg-
1 in soils (Adriano 2001) where Znenrichment through atmospheric depositions is considered reasonably high (Mertens and Smolders 2013). In addition, Zn levels in soil are very much related to soil parent materials, higher application rates of sewage sludge, mining and smelting activities (Nziguheba and Smolders 2008). Romans used various Zn alloys in 200 BC as a Cu alloy to make brass, while the production of Zn was started during the 20thcentury in Asia (Nriagu 1996). According to International Lead and Zinc Study Group (ILZSG), the production of Zn continuously increased with time and it was doubled every decade for the past twenty years with 11.3-106-ton total global production in 2007.
Zinc occurs in divalent form in soils. Dissolution, precipitation, sorption and solution complexation are the main reactions affecting soil solution Zn concentrations (Alloway 2009). Zinc is sorbed to oxyhydroxides of Fe, Al, and Mn and organic matter by inner sphere complexation or specific sorption. The rhizosphere pH is the most important factor for Zn uptake and absorption by plants (Mertens and Smolders 2013; Houben et al. 2014; Duraes et al. 2015). Zinc acquisition by plants is mainly increased through extended root growth in combination with the mycorrhizal associations, and chemical processes occurring in the rhizosphere such as release of root exudates, organic acids and phytosiderophores (Abbas et al. 2016).
As a plant nutrient, Zn is highly involved in numerous plant metabolic and physiological processes. It exists in divalent form in plants as well as in other biological systems (Kupper and Andresen 2016). It is the only metallic nutrient that is component of all the enzymes classes (Broadley et al. 2007). Zinc is a necessary structural component of Cu-Zn superoxide dismutase, RNA polymerase, alcohol dehydrogenase and DNA-binding proteins (Broadley et al. 2007). Zinc is also considered an important protective agent because it is involved in defence mechanisms of plant cells against free ROS radicles. Presence of Zn in plant is associated with the protein synthesis, ribosome’s structural integrity, carbohydrate metabolism, phosphate metabolisms as well as gene expression and regulation. Zinc plays structural and functional role in bio-membranes and also in their stability under oxidative damage by detoxifying ROS (Bell and Dell 2008).
Zinc can negatively affect plant metabolic reactions both at low (Zn deficiency) as well as at higher levels (Zn toxicity). Zinc deficiency is a global issue both in plants and humans now-a-days (Rehman et al. 2012). Globally, more than one third human population and agricultural area are facing the problem of Zn deficiency due to low Zn solubility and/or high Zn fixation (Alloway 2008; Nikolic et al. 2016). High lime (CaCO3) contents, nature of clay minerals, neutral to alkaline soil pH, high soil organic matter contents, higher concentrations of phosphorous or magnesium and prolonged periods of waterlogging are the main reasons associated with Zn deficiency (Oliver and Gregory 2015). Zinc deficiency could also induce photo-oxidative stress by enhancing the production of ROS. Human populations face Zn deficiency when they eat food products low in Zn which is grown on Zn deficient soils (Oliver and Gregory 2015). For instance, Zn concentration in Triticum aestivum grain is usually 30 mg Zn kg-1 while in Zn deficient Triticum aestivum grains; it
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may be 15 mg Zn kg-1 (Alloway 2008). Selection of crop varieties and Zn fertilization are common approaches to overcome Zn deficiency problems (Rehman et al. 2012; Fahad et al. 2015).
A high supply of Zn can rapidly induce its toxicity in non-tolerant plants causing inhibition of root elongation (Ruano et al. 1988). The toxicity of Zn normally occurs in plants grown in soils contaminated through anthropogenic activities like smelting, treated with sewage sludge and mining activities (Nannoni et al. 2016). Toxicity of Zn can cause deficiency of other nutrients such as Mg and Fe due to similar ionic radius (Sagardoy et al. 2009). At high levels, Zn can become phytotoxic, causing leaf chlorosis, decreased growth and nutritional disturbances. Zinc toxicity also imped normal ionic homeostatic systems by interfering with the phyto-absorption and transportation of essential nutrients, thereby causing disruption of metabolic reactions such as photosynthesis, transpiration and activity of metabolic enzymes. Under very high Zn stress, it may induce reactive oxygen species (ROS) production and thereby deteriorates proteins and forms complexes with RNA and DNA (Broadley et al. 2007).
7.2 GLOBAL USES OF ZINC
The industrial production of Zn started since the late 18th century with a continuous increase up till now. Almost one half of this industrially produced Zn is used for galvanizing other metals, such as iron, to prevent rusting (ATSDR 2015c; Mertens and Smolders 2013). Zinc is also used to produce die-castings, which are used for automobiles, hardware and electrical industries. It is used in many alloys such as nickel, silver, brass and aluminium solders. Zinc oxide is used for manufacturing many products such as rubber, paints, plastics, cosmetics, pharmaceuticals, soaps, inks, batteries, textiles and electrical equipment. Zinc sulfide is used in luminous paints, X-ray screens and fluorescent lights (ATSDR 2015c; Emsley 2001). Zinc production from mining increased by 25% from 10x106 to 13.4x106 metric tons from 2006 to 2015 (Fig. 7.1).
Figure 7.1 Annual world refinery production of zinc (USGS
2016).
9
10
11
12
13
14
10
3T
ho
usa
nd
s m
etr
ic t
on
s
Zinc
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7.3 ZINC LEVELS AND SOURCES IN SOILS
Zinc occurs naturally in the Earth crust as a component of various Zn-rich polymetallic ores and rocks. Zinc ores are formed by natural geological processes and are present around the globe. The normal range of Zn is 5–15% in economically important Zn ores mainly in the form of sulfides (wurzite or sphalerite). Other important Zn minerals are Zn silicates (hemimorphite) and zinc carbonates (smithsonite) (ATSDR 2015c; Mertens and Smolders 2013). Many countries such as China, Peru, USA, Canada and Iran are rich in Zn deposits. Soils of these Zn-rich locations have a considerably higher Zn concentration compared to the average soil Zn concentration. The basalt, basic igneous rocks, usually contains 110 mg Zn kg-1. This higher Zn concentration seems the result of isomorphic substitution of Zn for Fe2+ and Mg2+ in ferromagnesium minerals such as augite, hornblende and biotite (USGS 2016). On contrary, granite and gneiss are the examples of silica-rich igneous and metamorphic rocks; respectively those have lower Zn concentration (usually 40 mg Zn kg-1). The Zn concentrations of sedimentary rocks are dependent on the rocks from which these sedimentary rocks are formed. Sandstone and limestone usually have low Zn levels (20 mg Zn kg-1) because these originate from low Zn parent materials, whereas, clays may contain up to 100 mg Zn kg-1 and black shales which are Zn-rich and may contain much higher Zn (1,500 mg Zn kg-1) concentrations (USGS 2016).
Different values of natural background concentration of Zn have been reported as 60 mg kg-1 (Burt et al. 2003), 62 mg kg-1 (Kabata-Pendias and Pendias 2011), 65 mg kg-1 (Eriksson 2001), 68.1 mg kg-1 (FOREGS 2005), 73 mg kg-1 (Licht 2005) and 89 mg kg-1 (Takeda et al. 2004). Clayey soils usually have high Zn concentration compared to sandy soils because of their high native Zn content and more ability for adsorbing and retaining Zn. Baize (1997) concluded that, Zn concentration of sandy soils was 17 mg kg-1 whereas, clayey soils have up to132 mg kg-1 concentration in France. Cleven et al. (1993) observed that sandy soils contained 20–45 mg kg-1, peat soils have 55–140 mg kg-1 and clayey soils have 70–150 mg kg-1 Zn in the Netherlands.
Up to 2007, Asia was producing almost 55% of global Zn in smelters and refineries, whereas, America and Europe were producing 16% and 20%, respectively. In order to protect the environment, this global industrial Zn production was decreased from 3.4 ×106 to 2.7×106 tons from the early 1980s to the early 1990s (Nriagu1996). Between 1967 and 1990, atmospheric Zn emissions were decreased almost 2.5 times (Boutron et al. 1991). Similarly, between 1980 and 2000, there was a two time decrease in Zn emissions from a wastewater treatment facility in France (Grosbois et al. 2006). This decreased emission of Zn to soils and waters was mainly due to decrease in atmospheric SO2 which ultimately decreased the run-off of Zn from galvanizing facilities. Tyre debris along roadsides is another factor causing Zn contamination. Tyres usually have ZnO ranging from 1–2%, and Europe alone is responsible for release of more than 4,000 tons Zn annually via tyre debris (Blok 2005). Natural phenomena such as volcanic eruptions, forest fire and winds are collectively responsible for emitting 50 ×103 tons of Zn to the atmosphere annually and in 1983; these factors caused nearly 40% of the total annual Zn emissions (Nriagu 1989). Various human activities such as traffic, coal and fuel combustion caused rest of the 60% Zn emissions (EU 2009).
Zinc in the form of aerosols accounts for less than 1% of total atmospheric Zn. Levels of Zn in the air are usually in the range of 0.01–0.2 ug m-3. Industrial and urban areas have a bit higher Zn concentrations, but those are usually less than 1 ug m-3 (EU 2009). Rain and gravitational settling respectively, cause wet and dry depositions of atmospheric Zn onto soil, vegetation, water and other surfaces. Zinc is also added into soils through various agricultural activities such as the application of manures, sewage sludge, bio-solids and inorganic fertilizers. Presence of Zn in bedding materials such as straw, and addition of Zn supplements in the livestock diets are some other main sources of Zn. Zinc concentration varies among fertilizers; hence the type and amount of fertilizer determines the level of Zn in arable soils.
The concentration of Zn in inorganic fertilizers (except of Zn fertilizers) is very low. In Europe during 1999-2000, the input of Zn into agricultural soils by phosphatic fertilizers was only 43 g ha-1 year-1 (Nziguheba and Smolders 2008). Application of sewage sludge can add
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significant amount of Zn to soils. Zinc levels in sewage sludge are related to the source of sludge (either industry or domestic). Industry is responsible for adding more than half of the Zn to sewage sludge, whereas, the main sources of domestic Zn are the essential body care products. In EU, average Zn concentration in sludge is in the range of 142–2,000 mg Zn kg-1 on dry weight basis. There is a decreasing trend in the sewage sludge Zn concentrations due to awareness among communities (Andersen 2001). The annual removal of Zn with the harvested crops and leaching is generally less than the annual additions. Typically, the 25 cm top soil with 50 mg Zn kg-1 contains160 kg Zn ha-1 (Mertens and Smolders 2013).
7.4 SOIL PROPERTIES AND ZINC PHYTOAVAILABILITY
Zinc occurs in divalent form in soils. Dissolution, precipitation, sorption and solution complexation are the main reactions affecting soil solution Zn concentration. The solid-liquid distribution coefficient (Kd) is used for the quantification of Zn distribution from solid to solution phase. The coefficient Kd is defined as the ratio of the total Zn and soluble Zn concentration. The Kd value is not constant and it varies as the soil solution chemistry is changed. Sorption and mineral dissolution are the main controlling reactions for the solution Zn concentration. There are contradictory findings regarding the dominance of these reactions and their role in the modelling of Zn solubility. It has been predicted through models that precipitation of soluble Zn occurs when the total Zn in soils is > 200 mg Zn kg-1 (Mertens and Smolders 2013). Common precipitates of Zn include smithsonite (ZnCO3), zincite (ZnO) and hydrozincite (Zn5(CO3)2(OH)6 (Jacquat et al. 2008). Voegelin et al. (2008) reported Zn precipitation in 51 soils (average 5,000 mg Zn kg-1) which were irrigated with water having excess Zn rationally due to high soil Zn concentrations. Nevertheless, solubility of Zn and its subsequent adsorption increased as the total soil Zn increased, which suggest that the precipitation is not the only controlling factor of Zn solubility (Degryse et al. 2011).
Van Damme et al. (2010) have reported a precipitation-controlled Zn solubility in neutral soils, which contained > 4300 mg Zn kg-1. Moreover, the adsorbed (labile) Zn concentration was almost 1000 mg kg-1, regardless of the total Zn contents as expected from mineral solubility model (Van Damme et al. 2010). Sorption is the main reaction which controls the immediate Zn contents in the soil solution. Zinc is sorbed onto the surface of oxyhydroxides of Al, Fe and Mn, OM, forming inner sphere complexation or specific sorption. On the other hand, non-specific adsorption of Zn can take place through ion exchange process on surfaces of silicate clay minerals or OMfunctional groups. Weak selectivity, reversibility and low pH dependency are the main characteristics of these reactions. It has been predicted that these reactions are important for Zn sorption only when the total concentrations of Zn are quite higher than the background Zn levels.
Due to complexation of Zn with inorganic salts such as sulphates and organic ligands (citric acid and humic acids), the reactivity of Zn decreased and its solubility increased. Generally, the solubility of Zn decreases as the pH increases. Stephan et al. (2008) determined the Zn speciation of 66 soil samples and concluded that 88 % of the total soluble Zn was in the form of organic-Zn complexes whereas, only 8 % of total soluble Zn was found as free Zn2+ ions. The ZnOH+ is predominant Zn species when soil pH is > 7, but if the solution pH is > 9, Zn(OH)2 will be the dominant species.
Initially, there is a rapid increase in adsorption of Zn onto solid surfaces after its application to soils. It is assumed that there is an immediate equilibrium between labile and solution phases. Metal ions adsorb onto soil by initial fast reaction, a relatively slow reaction called ageing or fixation takes place which decreases its mobility (desorption) into solution phase. These ageing reactions rationally increase the value of Kd and the phytoavailability of Zn is decreased. Intra-particle diffusion of Zn2+ in amorphous oxy-hydroxides is responsible for these aging reactions. This diffusion process is dependent on the electrostatic attraction forces. The Zn fixation is usually increased with increasing pH as a result of an increase in reactive sites on oxide minerals (Buekers 2007).
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The importance of this beneficial aging reaction can be visualized by comparing more toxicity of Zn in freshly spiked soils to the field contaminated soils. From plant nutrition point of view, these fixation reactions are not considered beneficial as they decrease the Zn phytoavailability (Brennan 1990). In an incubation experiment, Brennan (1990) studied these ageing reactions of Zn in 54 soils having different chemical properties. It was concluded that the ageing or fixation was more at higher pH values.
7.4.1 Rhizosphere and Zinc Phytoavailability
Zinc acquisition by plant roots is increased through (a) root growth and the mycorrhizae increase the area for Zn absorption, and (b) chemical processes occurring in the rhizosphere induced by plant roots in the form of root acidic exudation and phytosiderophores, which decrease rhizosphere pH. The change in rhizosphere pH along with root exudates of low and high molecular weight compounds play very important role in Zn absorption by plants growing on Zn deficient soils. Rhizosphere acidification increased the absorption of P (Kirk and Saleque1995) as well as that of Zn (Duraes et al. 2015) in many crops. Rhizosphere pH is changed due to (a) the excretion of H+ and OH− or HCO3
− ions as a result of imbalance absorption of cations and anions by plant roots, (b) the release of CO2 by plant roots through respiration, and (c) the exudation of different low molecular weight organic acids. According to Marschner (1995), the form in which nitrogen (N) fertilizer is applied has great influence on the ratio of cation to anion absorption, and thus indirectly affects rhizosphere pH. Under anaerobic soil conditions, release of CO2 from plant roots causes the extrusion of high amounts of H+ ions and thus the availability of Zn for plants is also increased (Kirk and Bajita 1995). Under such conditions, NH4
+ is the predominant for N absorption and it results in the release of H+ ions, whereby reducing the pH of rhizosphere. Under aerobic Oryza sativa cultivation, N is mainly absorbed in the form NO3
– leading to the exudation of OH− ions in the rhizosphere, which increases the rhizosphere pH and ultimately, the availability of Zn is decreased. The change in rhizosphere pH due to change in N supply is considered as one of the main reasons for limited Zn availability to Oryza sativa under oxidized calcareous soil (Gao et al. 2006).
Lowland Oryza sativa also exudes low molecular weight organic compounds. It was recorded that with increasing bicarbonate level in the growth medium, the exudation of malate was increased (Yang et al. 2003). Likewise, Hoffland et al. (2006) concluded that under Zn deficiency, the exudation of citric acid was increased in low land Oryza sativa genotypes. In aerobic Oryza sativa, very low concentration (0.5 mM) of malate was excreted both in hydroponics and soil culture experiments under Zn deficiency, and it was further concluded that this minor concentration was not sufficient to solubilize appreciable amount of Zn (Gao et al. 2009).
According to Rose et al. (2011), Zn concentration should be > 100 mM in soil solution for significant amount of Zn solubilisation and uptake. Hence, for aerobic Oryza sativa, mechanisms other than low molecular weight organic acids exudation are perhaps also important for Zn mobilization. Phytosiderophores release from Hordeum vulgare (Suzuki et al. 2006) and Triticum aestivum (Cakmak et al. 1994) roots has been found to increase under Zn deficiency. As against of many graminaceous crops, Oryza sativa has been found to grow well under submerged soil conditions. The release of phytosiderophores is quite low in Oryza sativa compared to other crops (Takagi 1976). These phytosiderophores are thought necessary for Fe absorption by Oryza sativa under aerobic conditions (Inoue et al. 2009), and for that of Zn under submerged soil conditions (Arnold et al. 2010).
According to Suzuki et al. (2008), release of phytosiderophores is more obvious in Oryza sativa under iron deficiency compared to Zn deficiency. The exact mechanism for this differential release is not clear yet, but it might be related to difference in growing conditions i.e. soil or hydroponics, and the Oryza sativa genotypes used in the experiments. The direct comparison of root exudations between flooded and non-flooded Oryza sativa, and its role in Zn availability to Oryza sativa plants has not been reported up till now. It is however believed that release of phytosiderophores under aerobic conditions is more important than submerged conditions. This is
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because, under aerobic conditions, exudates are more concentrated at the root surface due to their less movement from roots. More research is needed to investigate the mechanisms of phytosiderophores release from Oryza sativa roots and their effectiveness in Zn uptake by Oryza sativa.
7.4.2 Microbial Activity and Zinc behaviour in Soils
The beneficial effects of Arbuscular mycorrhizae fungi (AMF) are mostly observed in plants growing on Zn deficient soils. These fungi help to increase Zn absorption in many crops such as Triticum aestivum (Ryan and Angus 2003), lowland Oryza sativa (Purakayastha and Chhonkar 2001), Solanum lycopersicum (Cavagnaro et al. 2010) and Cajanus cajan (Wellings et al. 1991). Cavagnaro (2008) concluded that Zn absorption in plants colonized by AMF is increased either by enhanced Zn absorption by AMF or through indirect effects of AMF via physiological, nutrional and morphological changes in plant roots. Although, mycorrhizal fungi can survive in submerged soil conditions (Purakayastha and Chhonkar 2001), their activity is increased many-fold under aerobic conditions. Hajiboland et al. (2009) inoculated Oryza sativa roots with AMF in pots and concluded that under flooded conditions, root colonization was 27%, whereas, under non-flooded conditions it was 43%. Gao et al. (2007) in an aerobic pot experiment with six Oryza sativa genotypes grown on Zn deficient soil, observed a relatively higher (28–57%) root colonization by AMF which significantly increased Zn uptake by Oryza sativa. It was further observed that AMF increased Zn uptake only in those Oryza sativa genotypes that have inherently low Zn absorption capacity. It indicates a beneficial relationship between AMF and inherent Zn absorption capacity, and it was correlated to root-induced rhizosphere processes such as exudation of Zn chelators. It was concluded that plants with lower Zn mobilizing capability from Zn deficient soils usually show high response to mycorrhizal fungi. Recently, Rashid et al. (2016) entailed that nano-particles of ZnO are toxic to soil microbes and affect their functions such as nitrogen and carbon mineralization in soils.
7.5 SOIL-PLANT TRANSFER OF ZINC
7.5.1 Molecular Understanding of Zinc Absorption by Plants
Plants also need metals such as Ni, Zn, and Cu for better growth. However, absorption of these metals is not strictly selective. Plant Zn and P absorption increased when applied as Zn phosphate carbonate. Changes in the rhizosphere are of particular concern for Zn absorption from soils. Zinc phytoavailability is strongly inhibited due to elevated bicarbonate concentration, high soil pH and decomposition oxidation of OM. Among soil factors, pH has the most remarkable influence on Zn phytoavailability. Plant toxicity of Zn in Zn contaminated soils can be decreased by liming, thereby decreasing the Zn absorption of plant.
Metals are transported into shoots through their bulk flow in xylem tissues. Metals are transported from root symplast into xylem apoplast (Grennan 2009). There may be accumulation of metals including Zn in epidermis and trichomes at tissue level (Grennan 2009). Plants possess homeostasis to maintain adequate concentration of essential metal ions in various compartments and limit the damage by prolonged exposure of non-essential metal ions. Transporters of various families such as cation diffusion facilitators (CDF), natural resistance-associated macrophage (NRAMP), ZIP (ZRT, IRT-like protein) and ATP-binding cassettes (ABC) are implicated by the elemental cationic distribution in plants (Clemens 2001). The Zn is a cofactor of many enzymes like RNA polymerases, alcohol dehydrogenase, superoxide dismutase, carbonic anhydrase, and is involved in metabolism of carbohydrates, protein synthesis, nucleic acids and lipids (Cakmak 2000). The main constituents of metal homeostasis are its transport, chelation and sequestration processes (Clemens 2001). Iron and Zn uptake are mediated by transporters of ZIP family. Positively surface charged substrates are translocated across membranes by transporters like ATPases which get energy from exergonic ATP hydrolysis reaction. The HMA1-4 and HMA5-8 are assumed to transport Zn, Cd, Pb, Co and Cu/Ag respectively. Members of subgroup Zn, Cd, Pb, Co have only been identified in prokaryotes. According to Verret et al. (2004), over-
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expression of AtHMA4 in plants helps improve tolerance to Zn, Co and Cd. In shoots, Zn or Cd were translocated and sequestrated into cell walls to a greater extent compared with that for control when exposed to high levels of Zn or Cd. In roots, it enhanced metal loading into the xylem.
7.5.2 Competition between Zinc and other Cations for Absorption
Zinc is considered the most yields limiting micronutrient in some parts of the world (Sharma 2013). Availability and phytoabsorption of Zn is inhibited by its immobile nature even if high amount of soil Zn is present. Zinc plant absorption depends on its solubility which varies with the form and composition of the fertilizers. Plant availability is altered with soil and plant factors. Efficient use of Zn containing fertilizers demands a thorough understanding of Zn behaviour in soils. The Zn deficiency in plants can be overcome by agronomic biofortification through applying Zn containing fertilizers (Watts-Williams et al. 2014). The most common and typical Zn fertilizer for agriculture is hydrated zinc sulphate (ZnSO4.7H2O) or Zinc oxide (ZnO). Zinc carrying fertilizers such as Zn-NPK or Zn-coated superphosphate or urea are also used now-a-days in agriculture. Zinc absorption by plants is inhibited at high pH, oxides and hydroxides of Al and Fe and H+ generated by phosphate salts (Robson 1993). Zinc level in soils also affected the Cd absorption by plants (Hussain et al. 2016).
Plant absorption of Zn considerably depends on phytosiderphores, Zn transporters and mycorrhizal associations. Watts-Williams et al. (2014) performed glass house experiments to investigate the influence of simultaneous application of phosphorus (P) and Zn fertilizers on their availability with and without Arbuscular Mycorrhizae and found antagonistic effects on plants absorption of Zn. Phytoavailability of Zn and P was improved under application of Zn phosphate carbonate. They reported that non-mycorrhizal plants took less P than mycorrhizal with reverse behaviour by Zn absorption. Enhanced P uptake in plants improves plant growth leading to Zn deficiency due to growth dilution effect caused by growth improvement. Zinc mobility is decreased under sufficient P nutritional status. Conditions of bulk soils and rhizosphere vary in different aspects due to root exudates and microbial activities. Plant responses to the deficiency of mineral nutrients, absorption of certain cations/anions and excretion of organic acids induce rhizosphere acidification, i.e. P deficiency (Ohwaki and Hirata 1992). Graminaceous plant species release non-protein organic amino acids (phytosiderophores) due to Fe deficiency which form chelates of Fe, Cu and Zn. Similarly, Zn deficiency induces rhizosphere modifications. However, application of NO3-N in dicots results in rhizosphere acidification due to increased anion/cation uptake ratio (Cakmak and Marschner 1990). Root induced changes and presence of microorganisms change the solubility and mobility of nutrients. Supply of the mineral nutrients of low mobility and metals, e.g. Cu and Zn, can be enhanced by vesicular-arbuscular mycorrhizae (Watts-Williams et al. 2014).
7.5.3 Zinc Sequestration into Plant Roots
Phytoavailability of Zn for plant absorption is not only affected by soil pool size of Zn but also by its concentration in the solution and its transport to root surfaces. The distance between the source and sink of Zn is actually related to the surface area and development of roots (Alloway and Trevors 2013). Diffusion is the main driving force for Zn supply to plants in the zone around roots which does not extend beyond root hairs. Root growth, root exudates, morphology and surface area are important parameters affecting Zn accessibility and phytoavailability. There are several environmental and soil factors which affect the correlations between the extractable Zn and its phytoavailability (Lindsay and Cox 1985). These factors must be considered to improve the correlations between extractable Zn and plant availability. In calcareous soils and liming of acidic soils, Zn diffusion to plant roots may be about 50 times lower than in acidic soils (Moraghan and Mascagni 1991). Zinc contents are conspicuously declined in plants with liming acid soils. In fact, liming acid soils is an effective procedure to decease Zn availability and its toxicity in plants grown on Zn contaminated sites. The increased soil temperature improves Zn supply into soil solution, thereby, desorption of the fixed Zn. Soil microbes also increase Zn in soil solution due to
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mobilization of insoluble Zn forms (Mubiana and Blust 2007). Soil water contents and Zn contents in plants are often poorly correlated with the extractable Zn contents in soils. Solubility of Zn is generally controlled by sorption processes. Zinc deficient soils are wide spread than the Zn contaminated soils all over the world. Fixation is important in decreasing Zn availability to plants. Deng et al. (2004) studied accumulation pattern of Pb, Zn, Cu and Cd in 12 wetland plant species at contaminated sites in China. Root tissues showed maximum accumulation of metals in wetland plants suggesting exclusion strategy for metal tolerance.
7.5.4 Zinc Translocation into Plant Shoots
Vacuolar and apoplastic compartments in shoots are potential metal sequestration sections. Availability of metals is influenced by several environmental and biological factors. The solution chemistry and physical kinetics vary with temperature and affect metal availability to plants but its impact on metals bioaccumulation is still not adequately understood. There are significant variations in susceptibility to Zn deficiency in plants among cereal species and cultivars of a species (Erenoglu et al. 2002). Ekiz et al. (1998) studied the sequential behaviour of cereals susceptibility to Zn deficiency. Over-expressions of AtHMA4 probably cause an improved cellular metal efflux towards shoot and root apoplasts (Verret et al. 2004). Some species of wetland plants could accumulate relatively high metals including Zn concentrations in shoots (Deng et al. 2004).
7.5.5 Foliar Absorption of Zinc
Foliar application of Zn fertilizers is significant in fortifying crop yields and intake for human consumption. There is limited information regarding transport mechanisms of Zn from leaves to other plant organs. Effectiveness of foliar applied nutrients is determined by the rate at which nutrients are absorbed by leaves and transported to other plant parts. Cuticle and epidermal cells are considered critical for initial penetration of the sprayed nutrients while stomata have minor impact on their absorption (Fernandez and Brown 2013).
After absorption of foliar applied Zn, phloem is undoubtedly involved in transporting it to other plant parts after subsequent loading to vascular tissues (Fernandez and Brown 2013). Recently, Du (2015) reviewed absorption of foliar applied Zn hydroxide nitrate or zinc nitrate in Solanum lycopersicum leaves using synchrotron-based X-ray fluorescence microscopy. In case of decreased mobility, plants may show symptoms of malnutrition. The concentration of Zn in the underlying tissues in Solanum lycopersicum increased 600 folds. As a whole, Zn moved to lower veins approximately 2-10 folds compared to adjacent veins of foliar applied Zn (Du 2015). Absorption of the applied Zn is affected with variations in the densities of trichomes, stomata, leaf age and nutrient deficiency (Schilmiller et al. 2008). Erenoglu et al. (2002) studied the role of Zn nutritional status of plants on its mobility by phloem in Triticum aestivum and Triticum durum differing in foliar applied Zn absorption, transport and effectiveness. In this short experiment, all the cultivars showed similar Zn absorption regardless of leaf age and nutritional status for applied ZnSO4.
7.6 ROLES OF ZINC IN PLANTS
Zinc as a plant nutrient is the 2nd most abundant transition metal after Fe and is involved in numerous plant metabolic and physiological processes. It exists in divalent form in plants as well as in other biological systems. It is the only metallic nutrient that is component of all the enzymes classes (lyases, oxidoreductases, hydrolases, isomerases, ligases and transferases) having both structural and catalytic role (Kupper and Andresen2016). It also plays structural and functional role in bio-membranes and their stability under oxidative stress by detoxifying ROS (Bell and Dell 2008). Zinc deficiency causes a decrease in protein synthesis via decreased RNA levels and interruption in DNA transcription, and gene regulation.
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7.6.1 Role of Zinc in Enzymes
As explained above, Zn is the only metallic element that is part of all the six enzymes classes. There are four types of Zn-binding sites in these enzymes namely: structural, catalytic, protein interface and co-catalytic responsible for the activity of these enzymes (Sousa et al. 2009). A catalytic Zn ion is located at active site of enzymes, Zn ions are bound to one water molecule and three protein ligands, e.g. carbonic anhydrase. In structural Zn-sites, it is bound by four cysteine residues. Structural Zn sites are responsible for maintaining the structure of enzymes such as proteins and alcohol dehydrogenase involved in gene expression and DNA replication. In enzymes containing co-catalytic Zn-sites, two or more Zn atoms are bound by histidine and aspartic acid mostly (Auld 2009). Portion of Zn-binding ligands in enzymes is in the order of: histidine (28%), cysteine (23%), water (15%), aspartic acid (13%), glutamic acid (11%) and others (10%) (Sousa et al. 2009). Two types of hypothesis are presented about the catalytic action of Zn in enzymes (Vallee 1983); (i) Zinc-carbonyl hypothesis: “The substrates bind directly to Zn to form an enzyme-substrate complex and displace metal-bound H2O molecules in the process; i.e. Zn activates the electrophile (aldolase, peptidase)”. (ii) Zinc-hydroxide hypothesis: “The substrate does not bind directly to the Zn but mediates its function through the metal-bound water molecule”. The resultant metal-bound hydroxide ions can then attack the substrate; Zn also helps activate the nucleophile (carbonic anhydrase, citric acid).
Apart from the above two mechanisms, an integrated approach can also operate including both Zn-carbonyl and Zn-hydroxide hypothesis, Zinc acts as polarizing the substrates and activating water molecules. Whatever, the mechanisms of catalytic action of Zn in enzymes are, the involvement of Zn in activity of various enzymes suggests that it has significant role in plant metabolism. Generally, the metabolism of proteins, carbohydrates, auxin and reproductive processes are severely affected under Zn-deficiency.
a. Alcohol dehydrogenase
Alcohol dehydrogenase is exceptional among Zn-containing enzymes since have two Zn atoms, while most of the other enzymes contain one Zn atom. One of the Zn atoms plays structural role while another catalytic role (Auld and Bergman 2008). Catalytic Zn-sites are occupied by two cysteine, a water and one histidine molecules. Structural Zn-binding sites are bound by four cysteines. Alcohol dehydrogenase catalyses the conversion of acetaldehyde to ethanol.
Under anaerobic conditions, ethanol formation occurs in meristematic tissues, like root apices. Plants deficient in Zn show decreased alcohol dehydrogenase activity and its effect on plant metabolism under aerobic conditions is not clear yet. However, under submerged soil conditions, alcohol dehydrogenase activity in plant is shown to be twice in Zn-sufficient than that in Zn-deficient plants.
b. Carbonic anhydrase
Carbonic anhydrase (CA) have one Zn atom to catalyse the hydration of CO2 to convert it into HCO3
- and H+. This enzyme is localized in cytoplasm and chloroplast. Dicotyledonous plants have CA in six subunits and six atoms of Zn per molecule having molecular weight of 180 kDa (Sandmann and Boger 1983).
The role of carbonic anhydrase differs among C3 and C4 plants, especially in chloroplast and also between bundle sheath and chloroplast in C4 plants. In C3 plants, it assists the diffusion of CO2 to the site of carboxylation, where it is reduced by rubisco. In C4 plants it reduces CO2 to HCO3
- to be used by phosphoenolpyruvate-carboxylase (Badger and Price 1994). Zinc-nutritional status does not directly affect the photosynthetic CO2 assimilation by carbonic anhydrase activity in C3 plants. Zinc deficiency in Oryza sativa plants showed decreased expression of mRNA for CA up to 13% suggesting a decrease in CA activity (Sasaki et al. 1998). High Zn deficiency completely inhibits CA activity. However, with low CA activity maximum net photosynthesis can occur. The C4 plants differ from C3 plants in the mechanism of CA activity (Hatch and Burnell
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1990). A higher activity of CA is essential in the mesophyll and chloroplasts to form substrate for phosphor-enolpyruvate-carboxylase (HCO3
-), which is then converted to C4 compounds, e.g. malate. In this process, CO2 is released and acts as a substrate for RuBP-carboxylase. In C4 plants, in vivo activity of CA is enough to check the conversion of CO2 to HCO3
- from limiting photosynthesis (Hatch and Burnell 1990). Hence, deficiency of Zn may have a sever influence on photosynthetic rate in C4 than that in C3 plants (Burnell et al. 1990).
c. CuZn-superoxide dismutase
The CuZn-superoxide dismutase (CuZn-SOD) is the most abundant isoform of SOD in plant cells. It has a molecular weight of 32.5 kDa and is associated in the scavenging of superoxide radicals (O2
.-), particularly under abiotic stresses. In CuZn-SOD, Zn provides the structural stability and is complexed with two histidine and one aspartate molecules (Abreu and Cabelli 2010) while, Cu performs the catalytic functions. Literature suggests that Zn is an important structural component of CuZn-SOD as its deficiency negatively affects the activity of CuZn-SOD and resupply enhanced it (Hacisalihoglu et al. 2003). Thus, the activity of CuZn-SOD is very good indicator plant tolerance under Zn deficiency (Hacisalihoglu et al. 2003).
The activity of SOD is critical under abiotic stresses as it is the determining factor regarding the tolerance of plants. Zinc deficiency under such abiotic conditions might cause failure of plants because of the decreased SOD activity (Cakmak and Marschner 1988). Accordingly, over-expression of activity of CuZn-superoxide dismutase in genetically engineered plants might enhance tolerance against numerous abiotic stresses (Kim et al. 2010).
d. Other Zn-containing enzymes
There are numerous other enzymes those require Zn both as structural and catalytic components for their functioning such as, phospholipase and alkaline phosphatise. Both these enzymes have three atoms of Zn, at least one of them has catalytic activity (Coleman 1998). Carboxypeptidase: having one Zn atom with catalytic function causing hydrolysis of peptide cleavages. The RNA-polymerase having two Zn atoms becomes inactive by removal of Zn (Falchuk et al. 1977). Activity of RNA polymerase has been extensively studied in relation to Zn nutrition in different species including animals, bacteria and humans. However, very little information is available in higher plants.
7.6.2 Zinc and Protein Synthesis
Zinc deficiency causes a decrease both in the rate of protein synthesis from amino acids and concentration. Whereas, re-supplying Zn to deficient plants, start protein synthesis immediately. Apart from this, Zn has other roles in protein metabolism. Zinc being structural part of ribosomes provides structural integrity to these molecules. Absence of Zn disintegrates ribosomes. However, restoring the Zn supply reconstitutes them. Zinc-deficiency immediately causes a sudden increase in the enzymatic activities which are involved in protein synthesis, like ribonuclease, ATP-ase and glutamic dehydrogenase leading to decrease in RNA and protein synthesis. In Oryza sativa, disintegration of soluble fraction of ribosomes (80S) takes place at Zn concentration < 100 µg g-1 of dry weight, whereas, in Nicotiana tabacum the concentration was 70 µg for 80S ribosomes and 50 µg for decrease in protein contents (Obata and Umebayashi 1988). High amino acid accumulation and low protein concentration by Zn deficiency is not only the result of decreased transcription and translocation but also at elevated rates of RNA degradation, because of high activity of RNA-ase under deficiency of Zn (Sharma et al. 1982). All this discussion shows that an inverse relationship exists between activity of RNA-ase and Zn-supply and between RNA-ase and plant protein contents.
7.6.3 Zinc and Membrane Stability
Zinc is very important for maintaining the membrane stability by lipid peroxidation caused by ROS under abiotic stress conditions. It protects membranes from oxidative damage by binding
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to sulfhydryl and phospholipid groups in membranes or by forming tetrahedral complexes with cysteine polypeptide chains. Since, Zn is a functional component of an important enzyme CuZn-SOD which controls the formation of cytotoxic ROS by causing interference in the oxidation of nicotinamide adenine dinucleotide phosphate (NADPH) and also by neutralizing superoxide ion (O2
·-) (Cakmak and Marschner 1988). Zinc deficiency causes an increase in plasma membrane permeability as shown by the outflow of low molecular weight solutes and phospholipid contents (Cakmak and Marschner 1988). Membrane integrity tends to restore immediately after re-supplying the Zn as early as after 12 hours. Increase in membrane permeability under Zn deficiency may also be attributable to higher rates of O2
·- formation due to elevated activity of NADPH-dependent O2
·- generating oxidase (Cakmak and Marschner 1988). This higher activity of O2
·- generating oxidase is indicative of the fact that Zn might have direct role in controllig enzyme activity or an indirect result of changes in structure and composition of membranes.
The obvious Zn-deficiency symptoms in plants (necrosis, increase in membrane permeability, chlorosis, and inhibited shoot growth), all are due to oxidative stress caused by excess generation of ROS beyond plants scavenging capacity and inefficient detoxification due to Zn deficiency. The antioxidant enzymes like CAT, POX, and APX involved in the scavenging of H2O2 also show marked decrease in their activity due to Zn deficiency which can be due to decreased synthesis of protein (Cakmak 2000). This induces an accumulation of ROS causing lipid peroxidation by Zn-deficiency and increasing membrane permeability (Chen et al. 2008). All these events are described in schematic diagram Fig. 7.2).
7.7 ZINC DEFICIENCY
Approximately 30% of the world agricultural lands are Zn deficient and one third population of the world is suffering from Zn deficiency (Nikolic et al. 2016). The main reason for Zn deficiency is its low bioavailability and fixation in soils. Total soil Zn concentration is not regarded as a good estimate for plant available Zn. The main reasons of Zn deficiency are: high lime (CaCO3) contents, alkaline soil pH, nature of clay contents, high soil organic matter, higher concentrations of P or Mg and prolonged periods of waterlogging (Oliver and Gregory. 2015). Sorption of Zn onto Al and Fe oxyhydroxides is increased when higher doses of P-fertilizers are applied.
Zinc deficiency is common in plants in calcareous and highly weathered soils (Oliver and Gregory 2015). In calcareous soils, Zn deficiency is normally linked with deficiency of Fe called “lime chlorosis”, due to high pH in calcareous soils that causes adsorption of Zn onto clay minerals or CaCO3 instead of forming sparingly soluble compounds of Zn i.e. ZnCO3 or Zn(OH)2. Also, the high concentration of HCO3
- strongly inhibits absorption and translocation of Zn, similar to the effect on Fe.
In dicotyledonous plants, typical visible symptoms of Zn-deficiency include “resetting,” i.e. limited plant growth because of shortening of intermodal distance and a significant decrease of leaf size. Severe Zn-deficiency causes a condition called “die-back” (dying of shoot apices). These symptoms are often accompanied by chlorosis which might be highly contrasting and diffusive (mottle leaf). Plants exposed to high light intensity, drought, and heat stress show high susceptibility to low Zn supply than plants grown in normal conditions (Peck and McDonald 2010). Shoot growth is normally more sensitive to Zn-deficiency than root growth, and root growth may get increased by conceding shoot growth (Cakmak et al. 1996). The Zn deficiency enhances the seceretion of low molecular weight solutes by roots. The exudates include sugars, amino acids and phenolics in dicotyledonous species, whereas it includes phytosiderophores in graminaceous species (Cakmak et al. 1994). The critical Zn-deficiency symptoms in leaves are observed below 15-20 µg g-1dry weight. Zinc deficiency causes more decrease in grain or seed yield than that in total dry weight due to decreased pollen fertility
Low Zn intake in humans can cause Zn deficiency when they are grown on Zn deficient soils (Hefferon 2015; Nikolic et al. 2016). For instance, Zn concentration in Triticum aestivum grains is usually 30 mg Zn kg-1 and in Zn deficient Triticum aestivum grains it may be 15 mg Zn
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kg-1 (Alloway 2008). The level of recommended daily Zn intake necessary for human health is quite ambiguous. The Reference Daily Intake of Zn (RDI) is 15 mg for adults while it may vary from 4 to 17 mg Zn for youngsters as suggested by WHO/FAO/IAEA (WHO 2004). Zinc requirements for pregnant women are higher compared to adult ones. Meat and dairy products are Zn rich diets. Poor farming communities often are Zn deficient due to more consumption of cereal grains compared to meat and dairy products (Oliver and Gregory 2015). A daily intake of 500 g dairy products and 250 g meat may contribute more than 20 mg Zn. On the other hand, only Triticum aestivum provides more than 70% of the daily energy intake in many rural areas of Central Asia and Middle East countries (Cakmak 2008). A daily intake of 400 g tubers or Oryza sativa by poor farmers provides as low as 10 mg Zn per day.
7.8 ZINC TOXICITY
7.8.1 Zinc Toxicity to Plant Growth
Although, Zn is essential micronutrients for plants, yet it becomes toxic to plants at higher concentrations (Ann. Table 6). Usually, Zn concentration in a healthy plant is 60 mg Zn kg-1 dry weight (Davis and Beckett 1978). If this concentration is increased up to 100–500 mg Zn kg-1 dry weight, it results in 10% decrease in plant yield. A high supply of Zn can rapidly induce its toxicity in non-tolerant plants causing inhibition of root elongation (Kupper and Andresen 2016). Zinc toxicity can induce deficiency of other nutrients such as Mg and Fe due to similar ionic radius of these nutrients (Sagardoy et al. 2009). This Fe and Mg induced deficiency can cause chlorosis and/or necrosis in young leaves. Zinc toxicity can also induce Mn deficiency (Ruano et al. 1987). Zinc toxicity tends to impair photosynthetic activities at different steps by a number of ways such as suppressing the activity of Rubisco via competition of Zn with Mg. Zinc toxicity also limited photosynthesis in Beta vulgaris by decreasing the CO2 at Rubisco as a result of decreased stomatal and mesophyll conductance of CO2 by 70% and 44%, respectively (Sagardoy et al. 2009). Zinc toxicity can also decrease the efficiency of PSII via substituting Mn by Zn in thylakoid membranes (Van Assche and Clijsters 1986). Marcato-Romain (2009) reported Zn-induced genotoxicity in Vicia faba seedlings.
Zinc toxicity to plants and soil microorganisms can be easily determined by applying increasing levels of Zn solutions to uncontaminated soils (spiking). The assessment of Zn toxicity can be accomplished over a short time (hours to weeks) after Zn application to soils. Usually, Zn toxicity is determined under controlled conditions in laboratories, but the applicability of such results might be doubtful for real field conditions. Firstly, the composition of soil solution is changed by the addition of Zn salts. The adsorption of Zn2+ causes desorption of many other cations such as Ca2+ and Mg2+, and it is supposed that soil acidification may be partly responsible for Zn toxicity (Speir et al. 1999). Secondly, ionic strength of soil solution is increased due to anionic Zn ligands (chlorides and sulphates) which may cause osmotic effects (Stevens et al. 2003).
7.8.2 Zinc-Induced Oxidative Stress
Zinc metal itself plays a key role in cellular defence system against ROS. Therefore, Zn is considered an excellent defensive metal against oxidative damage to cellular components such as chlorophylls, membrane lipids, and –SH groups of protein. However, at high levels, Zn becomes toxic and induces negative impact on plant growth and development. Inside plant cells, Zn toxicity induces oxidative stress (Fig. 7.2, Table 7.1). The term ‘‘oxidative stress’’ refers to the increased production of reactive oxygen species (ROS) in cells due to toxicity of organic or inorganic contaminants (Shahid et al. 2014a). The ROS are usually unstable, chemically very reactive and short lived molecules. These species contain unpaired electrons in their valence shell. These ROS include: superoxide anion (O2
• −), singlet oxygen (½O2), hydrogen peroxide (H2O2), hydroxyl (HO•), peroxyl (RO2
•), alkoxyl (RO•) radicals, organic hydroperoxide (ROOH). The ROS are naturally produced inside plants as by-products of normal aerobic metabolisms in plant organelles such as chloroplasts, mitochondria and peroxisomes. The ROS play important
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roles in the metabolism of plants which include growth regulation, initiation of defence metabolism under stress, programmed cell death through signal transduction, fruit ripening and senescence, regulation of stomatal movement and alleviation of seed dormancy (Pourrut et al. 2011; Gupta et al. 2013). However, the steady-state level of ROS under normal physiological conditions is governed by balancing the generation of ROS with their elimination by scavenging systems.
Figure 7.2 . Entry of zinc into plant cell walls and its
detoxification mechanisms.
Zinc stress (under deficiency and toxicity) triggers significant production of ROS (Feigl et al. 2014). Redox-active metals such as Fe and Cu can generate ROS directly via Haber-Weiss Reaction, while non-redox metals (Cd, Pb, Hg and Zn) induce ROS generation indirectly by disrupting pro- and anti-oxidative balance (Duan et al. 2015). Zinc is transition metal and therefore induced oxidative damage by mediating ROS production such as O2
−, H2O2 and HO• radical. Zinc-induced oxidation stress causes lipid peroxidation, membrane damage and enzyme inactivation in cells (Michael and Krishnaswamy 2014; Arenas-Lago et al. 2016). Plants experience “oxidative stress” when ROS are not efficiently scavenged. Oxidative injuries caused by oxidative stress result in decreased plant development and growth (Ogawa and Iwabuchi 2001). Under high Zn application, Zn-induced increase in ROS generation has been reported in various plant species such as Triticum aestivum (Xu et al. 2014), Brassica juncea (Feigl et al.
Oxidative
Stress
Cell wall
ROS
absorption
CuZn-SOD
Phospholipase
Phosphatase
H2O2
O2·
Apoplastic,
vacuolar chambers Zn2+
Plant Zn2+ forms
Zn2+
Free ions, storage
metallo-proteins, low
molecular weight
complexes
Detoxification by
enzymes
Alcohol dehydrogenase,
Carbonic anhydrase
Chlorosis , desorption of H+, Ca2+
and Mg2+, damage to membrane
lipid, membrane damage
Zn induced oxidative
stress
Figure: Entry of Zinc in plant cell wall and its detoxification mechanism
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2014), Zea mays (Wang and Jin 2007), Pisum sativum (Malecka et al. 2012), Solanum melongena (Wu et al. 2015), Verbascum thapsus (Morina et al. 2010) and Myracrodruon urundeuva (Gomes et al. 2013).
7.9 ZINC DETOXIFICATION IN PLANTS
Plants have developed several mechanisms for detoxification against heavy metal exposure and accumulation to decrease their harmful effects mainly based on chelation and subcellular sequestration. Plants can detoxify metals by sequestering them into vacuoles, binding with various thiol compounds and pumping them out of plasma membrane. Exposure to Zn has been reported to stimulate enzymatic [APX, CAT, GPX, SOD, GR and non-enzymatic antioxidants (PCs, GSH and ASC).
7.9.1 Vacuolar Compartmentalization of Zinc
An active metabolic process produces chelating compounds when a heavy metal exceeds its threshold level inside cells. Specific peptides (PCs and MTs) chelate metals in the cytosol. Specific subcellular compartments serve for the sequestration of heavy metals after their chelating process. Many small molecules, like organic acids, amino acids and phosphate derivatives, also take part in metal chelation process inside plant cell (Rauser 1999). When intracellular concentration of metals including Zn in plants reaches a high level, plants export ions into the apoplast or to compartmentalize by using their efflux pumps. Vacuoles are the main sink for metal ions (Shahid et al. 2014a). Plant exposure to high levels of Zn induced AtMTP3 expression resulting in Zn sequestration into vacuoles (Peng et al. 2015). Gustin et al. (2009) showed MTP1-dependent enhanced sequestration of Zn into shoot vacuoles of Zn-hyperaccumulating Arabidopsis thaliana suggests dual roles in Zn accumulation and tolerance in plants.
7.9.2 Zinc Chelation by Phytochelatins
Phytochelatins (PCs) are heavy metal binding peptides. Plant exposure to toxic metals causes synthesis of PCs. Phytochelatins are produced in plant cells from reduced GSH by the activity of constitutive PC-synthase enzyme. Many studies have reported the role of PCs in heavy metals homeostasis and detoxification (Shahid et al. 2014a). Phytochelatins form metal-PC complexes and transport them into vacuole (Yadav 2010). In this way, PCs protect plants from the negative effect of toxic metals including Zn. Now-a-days, PCs have also been found associated in long-distance transportation of metals in the phloem. Barrameda-Medina et al. (2014) reported Zn-induced accumulation of PCs synthesis in Lactuca sativa. Tennstedt et al. (2008) showed that PC-deficient mutants (Arabidopsis thaliana) showed pronounced Zn2+ hypersensitivity.
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Table 7.1: Zinc-induced enhanced production of ROS in different plant species grown in
hydroponic
ROS Plant species Zn exposure level Duration Reference
H2O2, O2- Triticum aestivum 2 µM 20 days Xu et al. 2014
NO, O2- Brassica juncea 50, 150,300 µM 7 days Feigl et al. 2014
NO, O2- Brassica napus 50, 150, 300 µM 7 days Feigl et al. 2014
H2O2 Vigna unguiculata 5, 50 mg kg-1 14 days Michael and Krishnaswamy 2014
H2O2, O2- Zea maysφ 3, 9, 27, 81 mg kg−1 73 days Wang and Jin 2007
H2O2, O2- Zea mays 0.1, 10, 20µM 30 days Pandey et al. 2002
H2O2 Morus alba 1.0 µM 20 days Tewari et al. 2008
H2O2, O2- Pisum sativum 50 µM 4 days Malecka et al. 2012
H2O2, O2- Solanum melongena 10µM 7 days Wu et al. 2015
OH-, H2O2 Verbascum thapsus 1, 5 mM - Morina et al. 2010
H2O2 Myracrodruon urundeuvaφ 50, 80, 120,200mg kg−1
120 days Gomes et al. 2013
H2O2 Triticum aestivum 0.5, 1, 3 mM 6 days Li et al. 2013
H2O2 Cistus monspeliensis 500, 1000, 1500, 2000 µM
45 days Arenas-Lago et al 2016
O2-, OH-, H2O2, Ceratophyllum demersum 10, 50, 100, 200 µmol/L
7 days Aravind et al. 2009
H2O2 Triticum aestivum 3 mM 6 days Duan et al. 2015
O2 −; superoxide anion, HO•; hydroxyl, H2O2; hydrogen peroxide. Φ Grown in soil.
7.9.3 Zinc Chelation by Glutathione
Plants have established a very effective mechanism to combat with toxic effects of heavy metals including Zn. Glutathione (GSH) is considered the major source of non-protein thiols in plants, which plays key role in a number of biochemical and metabolic functions in almost all organisms including plants. Under metal stress conditions, plants can generate low molecular weight thiols that induce high binding affinity for heavy metals (Bricker et al. 2001). Among these
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defensive systems, ascorbate–glutathione (AsA–GSH) cycle is an important metabolic process in plants (Foyer and Noctor 2011; Shahid et al. 2014a). In this cycle, APX converts H2O2 into water. During this reaction, GSH donates electrons to dehydro-ascorbate reductase (DHAR) which reduces dehydroascorbic acid. The reduced glutathione (GSH) is simultaneously oxidized into glutathione disulfide (GSSG) by DHAR. In this way, AsA–GSH cycle maintains a proper redox environment. It is concluded that AsA–GSH cycle metabolism is linked with Zn-induced oxidative stress in plants (Barrameda-Medina et al. 2014). For example, Zn exposure to Lactuca sativa causes a significant increase in GSH and AsA, and abated GR activity (Barrameda-Medina et al. 2014). The AsA–GSH cycle is maintained by establishing high levels of GSH and AsA is associated with plant defence metabolism under metal stress (Shahid et al. 2014a). Wongkaew et al. (2015) reported that GSH in transgenic Arabidopsis considerably increased shoot tissues compared with wild type, which could be linked with high Zn contents in shoot tissues. Cuypers et al. (2001) recorded an increase in ascorbate oxidation and capacities constituting the AsA–GSH pathway in response to excess of Zn in Phaseolus vulgaris.
7.9.4 Antioxidants Enzymes and Zinc Toxicity
In order to defend Zn-induced oxidative damage, plants have developed mechanisms to scavenge ROS by synthesizing both enzymatic and non-enzymatic antioxidants (Table 7. 2). During ROS scavenging, antioxidants generally donate an electron to free radicals of ROS to form neutral and harmless end products such as water and oxygen. Superoxide dismutases (SOD) are key enzymes to catalyse the conversion of O2
− to H2O2. Zinc-mediated increase in the activity of antioxidant enzymes is reported for SOD in Vigna unguiculata, Morus alba, Zea mays, Jatropha curcas, Nicotiana tabacum and Pisum sativum (Michael and Krishnaswamy 2014; Tewari et al. 2008; Tavallali et al. 2010; Weisany et al. 2012; Tkalec et al. 2014; Luo et al. 2010; Malecka et al. 2012 and Hernandez-Viezcas et al. 2011). Other antioxidant enzymes (CAT, APX and POD) are involved in the removal of H2O2 (Shahid et al. 2014a). Zinc is also reported to decrease the activity of CAT in Morus alba, Zea mays, Pistacia vera, and Lycopersicon esculentum (Tavallali et al. 2010; Sbartai et al. 2012; Malecka et al. 2012). Zinc causes increase in POD in Morus alba, Zea mays, Triticum aestivum and Solanum lycopersicum (Li et al. 2013; Cherif et al. 2011; Tewari et al. 2008). Similarly, Zn stress induced GR activity in Zea mays, Nicotiana tabacum, Triticum aestivum and Solanum lycopersicum (Pandey et al. 2002; Tewari et al. 2008; Weisany et al. 2012).
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Table 7.2: Zinc-induced activation of antioxidant enzymes in different plant species grown in hydroponic
Enzyme Plant species Zn exposure level Duration References
SOD, POX, PPO Vigna unguiculata 5, 50 mg kg-1 14 days Michael and Krishnaswamy 2014
CAT, APX, POD, SOD Morus alba 1.0 µM 20 days Tewari et al. 2008
CAT, POD, SOD Zea maysφ 3.0, 9.0, 27.0, 81.0 mg kg−1
73 days Wang and Jin 2007
SOD, APX, GR Zea mays 0.1, 10, 20µM 30 days Pandey et al. 2002
LOX, APX, CAT, SOD Pistacia veraφ 5, 10,20 mg kg−1 100 days Tavallali et al. 2010
APX, POD, CAT Glycine maxφ 10 mg kg-1 42 days Weisany et al. 2012
CAT, APX, SOD, GR Nicotiana tabacum 25,50 µM 90 days Tkalec et al. 2014
CAT, POD, SOD Jatropha curcas 0. 25, 0.5, 1, 2, 3 mM 12 hrs Luo et al. 2010
CAT, SOD, GR Pisum sativum 50 µM 4 days Malecka et al. 2012
CAT, APOX Prosopis juliflora-velutina
0, 500, 1000, 2000, 4000 mg/l
15 days Hernandez-Viezcas et al. 2011
SOD, POD, CAT, MDA Brassica oleracea 1.0µM 16 days Hajiboland and Amirazad 2010
GSH, GSSG, GST Zea mays 10, 50, 75, 100µM 10 days Kleckerova et al. 2011
GPX,SOD, APX, CAT Oryza sativa 0.5, 5,50, 100 µM 14 days Asadi et al. 2012
AsA, NPT Glycine max 500, 1000, 2000 µmol/l 5 days Mishra and Prakash2010
SOD, CAT, APX, GR, GPX
Myracrodruon urundeuvaφ
50, 80, 120, 200mg kg−1
120 days Gomes et al. 2013
APX, SOD, POD, CAT, GR
Triticum aestivum 0.5, 1, 3 mM 6 days Li et al. 2013
APX, SOD, POD, CAT, GR
Solanum lycopersicum 10, 50, 100, 150 μmol/L
7 days Cherif et al. 2011
CAT, APX, GST Lycopersicon esculentum
50, 100, 250, 500 µM 7 days Sbartai et al. 2012
SOD; superoxide dismutase, APX; ascorbate peroxidase, GR; glutathione reductase, POD; Peroxidase, LOX; lipoxygenase, GPX; guaiacol peroxidase, CAT; catalase, AsA; ascorbic acid, GST; Glutathione S-transferase, AAO; ascorbic acid. Φ Grown in soil.
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7.10 CONCLUSIONS AND PERSPECTIVES
Zinc is abundant in the Earth crust and is present in almost all types of soils. In addition to many industrial applications, Zn is an essential plant nutrient. It is involved in many metabolic and physiological processes both in plants and animals including human beings. It exists in divalent form in plants as well as in other biological organisms/systems. Dissolution, precipitation, sorption and solution complexation are the main reactions affecting soil solution Zn concentrations. Rhizosphere pH is the most important factor for Zn absorption by plants. About one third human population and agricultural area of the world are Zn deficient. Generally, the metabolism of proteins, carbohydrates, auxin and reproductive processes are severely affected under Zn-deficiency. At high levels, Zn impairs various morphological, physiological and biochemical functions at cellular level. Zinc induces oxidative stress via over-production of ROS. Under Zn stress, plants activate numerous defence strategies to avoid its toxicity. These defensive approaches include Zn sequestration into vacuoles by forming complexes, binding Zn with GSH, PCs and amino acids. Moreover, numerous antioxidants are involved to combat increased production of Zn-mediated ROS in plants.
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200
ANNEXURES
Annexure Table 1: Harmful effects of arsenic on plant growth
Arsenic applied level
Duration Arsenic toxic effects on plant Plant
species Growth medium
Reference
134, 668 μM 8 days Root abnormality, decreased fresh biomass of leaves
Zea mays Hydroponics Duquesnoy et al. 2010
250, 350, 450 μM
12 days Decreased root growth, lateral ramifications in roots, anatomical alterations in roots
Brassicacea
e family Semi solid culture
Freitas-Silva et al. 2016
25, 50 μM 1,5 days Decreased root and leaf fresh weight, color change in root and leaves
Arabidopsis thaliana
Hydroponics Gupta et al. 2013
0.1, 0.3, 0.5 mM 30, 60 days
Decreased shoot length, decreased number of leaves
Brassica juncea
Soil Kanwar et al. 2015
2.5, 5.0, 10 μM 10 days
Inhibition in the growth of plants, electrolyte leakage, decreased cellular viability, decreased chlorophyll content
Phaseolus aureus
Hydroponics Malik et al. 2012
5, 10, 50 mg kg−1
70 days Decreased Zn content in shoot, decreased the chl and carotenoid contents
Trifolium pratense
Soil Mascher et al. 2002
25, 50, 75, 100 μM
7 days Stunted growth roots, chlorosis, decreased root dry weight, decreased chl content
Nasturtium officinale
Hydroponics Namdjoyan and Kermanian 2013
100 μM 1, 2 days Decreased root and shoot length, decreased root and shoot dry mass, chlorosis, cracked on root
Oryza sativa
Hydroponics Nath et al. 2014
5, 10, 20, 40, 80 μM
21 days Decreased root length, affected root and shoot growth
Cynara cardunculus
Hydroponics Sánchez-Pardo et al. 2015
5, 50 μM 7 days
Decreased root and shoot length, declined fresh weight of root and shoot, decreased Chl-a and b content, affected photosynthetic performance
Luffa acutangula
Hydroponics Singh et al. 2013
200, 500, 800 μM
90 days Decreased shoot and root growth Tamarix gallica
Soil Sghaier et al. 2015
50, 100, 500 μM 10 days Decreased germination, affected root and shoot growth
Oryza sativa
Hydroponics Shri et al. 2009
201
25, 50, 100 μM 20 days Decreased membrane stabilit, decreased photosynthetic pigments, poor germination
Hordeum vulgare
Sand Shukla et al. 2015
50,100 mg kg−1 120 days
Decreased dry and fresh weight of shoots, decreased photosynthetic rate, decreased Tran-spiration ratesand stomatal conductance
Myracrodruon urundeuva
Prepared test substrates
Gomes et al. 2014
6.6, 13. 2, 26.4, 52.8 μmol/l
3 days Membrane damaged Lactuca sativa
Hydroponics Gusman et al. 2013
5 mg/l 14 days Decreased fresh weight of roots and shoots
Nicotiana tabacum
Hydroponics Han et al. 2015
5, 25 μM 4 days Decreased shoot and root dry weight and length
Brassica juncea
Hydroponics Khan et al. 2009
20, 25, 50 μΜ 7 days Decreased root and shoot length Oryza sativa
Hydroponics Rai et al. 2011
2.5 mg/l 14 days Decreased biomass in root and shoot
Oryza sativa
Hydroponics Ren et al. 2014
25 μM 10 days Decreased shoot length and biomass
Oryza sativa
Hydroponics Tripath et al. 2013
5, 10, 15, 20, 25, 30, 35 mg kg−1
------ Affected stomata Vigna radiata
Soil Gupta andBhatnagar 2015
33.5, 67 μM 14days Seedling wilted, decreased shoot height and lower number of leaves
Brassica rapa
Hydroponics Shaibura and Kawai 2009
13.3 μΜ 7 days Inhibition of root growth Oryza sativa
Hydroponics Norton et al. 2008
202
Annexure Table 2: Harmful effects of lead on plant growth
Pb applied level
Durati
on Pb toxic effects on plant Plant Species
Growth Medium
Reference
15, 30 mg/l 30 days
Inhibition of photosynthesis and respiration, decreased plant growth, degeneration of cell organelles
Helianthus annuus
Hydroponics Mukhtar et al. 2010
10, 20 mg/l 16 hrs Reduction in plant growth
Alternanthera philoxeroides, Neyraudia
reynaudiana
Hydroponics Deng et al. 2006
15, 20, 30 mg/l
10 days
Loss of cell shape, decreased in the intercellular spaces, and shrinkage of root vascular bundle
Eichhornia crassipes
Hydroponics Baruah et al. 2012
40 µM 24 hrs Cell membrane damaged, photosynthetic rate decreased, induced stomatal closure
Salvinia minima
Hydroponics Leal-alvarado et al. 2016
0.005, 0.05 mg/l
14 days
Damaged root surface and cell walls
Raphanus sativus
Sand Hladun et al. 2015
50, 100, 200 mg/l
35 days
Damaged to the growth of plants, inhibited the formation of chlorophyll, inhibition of photosynthetic CO2 fixation
Brassica juncea
Hydroponics Niu et al. 2007
0. 25, 0.5, 0.75, 1.0, 1. 25 mM
1, 3, 5, 7 days
Effected chlorophylls biosynthesis, cell death, caused oxidative damage, chlorosis of leaves
Talinum triangulare
Hydroponics Kumar et al. 2012
100, 400, 600, 800, 1000 mg/l
10 days
Decreased plant biomass, caused oxidative damage, increased chlorophyll a, chlorophyll b contents
Sesbania grandiflora
Hydroponics Malar et al. 2014
0.5 mM 4 days Reduction in plant growth, reduction in their root calcium, zinc and copper contents
Lathyrus sativus
Hydroponics Brunet et al. 2008
100 µM 21 days
Effected photosynthetic system, chlorophyll a contents decreased, decreased Root Growth
Hirschfeldia incana
Hydroponics Fahr et al. 2015
10 mg/l 10 days
Affected plants root and shoots growth
Carex cinerascens
Hydroponics Liu et al. 2015
1 µM 7 days Root Length Growth decreased Matthiola flavida
Hydroponics Mohtadi et al. 2013
203
10 µmol/dm3 32 days
Influenced the transpiration, weight of roots decreased
Cucumis sativus
Hydroponics Varga et al. 1999
50, 100 µM 42 days
Inhibited various plant growth attributes, damaged to cell and sub-cellular organelles, reduction in protein synthesis
Brassica napus Hydroponics Shakoor et al. 2014
10 µM 30 days
Inhibition of growth, transpiration and Chl biosynthesis
Elymus elongatus
Hydroponics Sipos et al. 2013
25, 50, 100, 200 µM
5 days Effect on MDA and soluble protein Potamogeton crispus
Hydroponics Qiao et al. 2014
1.5, 3,15 mM 25days Reduction in plant growth, decreased in the contents of chlorophylls a and b and proline
Triticum aestivum
Hydroponics Lamhamdi et al. 2011
0-200 µM 1-7 days
Decreased root and shoot fresh biomass, decreases in photosynthetic pigments
Zea mays Hydroponics Gupta et al. 2009
40 µM 5 days Altered the metabolism and cell physiology
Salvinia minima
Hydroponics Estrella-Gómez et al. 2009
10, 25, 50 µmol/L
40 days
Decreased in leaves chlorophyll contents
Thalassia hemprichii
Hydroponics Purnama et al. 2015
1.5, 3, 15 mM
30 days
Metabolic changes, photosynthetic capacity, inhibition of plant growth.
Triticum aestivum
Hydroponics Lamhamdi et al. 2013
1.5, 3, 15 mM
30 days
Decreased soluble protein content, reduction in plant growth
Spinacia oleracea
Hydroponics Lamhamdi et al. 2013
100, 400, 800, 1500 µM
42 days
Impeded development of plants, interrupt activities of essential enzymes, photosynthetic processes and cells water usage
Cynara scolymus
Hydroponics Karimi et al. 2012
2.42, 4.83 mM
10 days
Decreased in amino acid, sugar, total phenolic and flavonoid contents
Spinacea oleracea
Hydroponics Khan et al. 2016
204
Annexure Table 3: Harmful effects of mercury on plant growth
Hg applied level Duration Hg toxic effects on plant Plant Species Growth Medium
Reference
1,2.5 mg/l 20 days Root and Shoot biomass decreased Oryza sativa Hydroponics Du et al. 2005
0 to 500 µm 10, 15 days
Decreased chlorophyll content, reduction in shoot and root length, elevated levels of lipid peroxides, protein oxidation
Cucumis sativus
Hydroponics Cargnelutti et al. 2006
0.05,0.1, 1 mg/l 28 days Growth inhibition of both roots and shoots
Atriplex codonocarpa
Hydroponics Lomonte et al. 2010
6, 30 µm 7 days Growth inhibition Zea mays Hydroponics Rellan-Alvarez et al. 2006
3, 10, 30 μM 1 day Higher cell death rate Medicago sativa
Hydroponics Ortega-Villasante et al. 2007
5, 10, 25, 50, 100 μM
14 days
Reduction in biomass and leaf water content, changed leaf cellular structure, decreased in palisade and spongy parenchyma cells
Brassica juncea
Hydroponics Shiyab et al. 2009
2.5, 5, 10, 25 μM 8 hrs Decreased root and shoot growth, chlorophyll contents, and total soluble proteins
Triticum aestivum
Hydroponics Sahu et al. 2012
10-50 μM 20 days Reduction in plant growth Solanum lycopersicum
Hydroponics Cho and Park2000
1–2 mg/L _ Decreased the growth Pisum sativum Hydroponics Beauford et al. 1977
1.0 μg/ml 10 days Inhibition of root and shoot growth Nicotiana Hydroponics Suszcynsky and Shann 1995
20 μg/l 15 days Inhibition of growth Suaeda salsa Hydroponics Wu et al. 2012
1, 3, 57 mM 10 days Decreased high percentage in seed germination, root and shoot length and dry weight of seedlings
Albizia lebbeck Hydroponics Iqbal et al. 2014
1, 5, 10, 20, 50 mg/l
10 days Chlorosis, browning of leaf tip, reduction in growth, stunting of seedlings
Cajanus cajan Hydroponics Patnaik and Mohanty 2013
6, 30 µM 7 days Reduction in root and shoot fresh weight
Zea mays Hydroponics Rellan-Alvarez et al. 2006
205
1, 5, 10, 20, 40 µM
0, 6, 12, 24, 48, 72 hrs
Reduction in root length, loss of plasma membrane integrity
Medicago sativa
Hydroponics Zhou et al. 2007
1, 5, 10, 20, 40 µM
0, 6, 12, 24, 48, 72 hrs
Oxidative damage to membrane lipids,over-generation of GSH in leaves
Medicago sativa
Hydroponics Zhou et al. 2008
0. 2 mM 10 days Chlorsis, reduction in chlorophyll, disrupt the metabolic balance and inactivate the antioxidant pool
Oryza sativa Hydroponics Wang et al. 2009
12.5, 100 µM 5 days Inhibited plant growth and development, affect cellular-level functions
Oryza sativa Hydroponics Chen et al. 2012
10, 25, 40, 50 μM 21 days Growth inhibition, chromosomal alterations
Sesbania drummondii
Hydroponics Israr and Sahi 2006
0.1, 0.5, 1.0, 2.5, 5 mg/l
20 days Reduction in root and shoot biomass Oryza sativa Hydroponics Du et al. 2005
1, 3, 5, 7 mM 7 days Decreased seed germination, reduction in seedling and root length
Vigna radiata Hydroponics Iqbal and Siddiqui 2015
1, 2, 3, 4, 5 mg/l 21 days Reduction in seed germination, seedling, root and shoot length
Arachies hypogiea
Hydroponics Abraham and Damodharam 2012
25, 50, 75, 100 μM
10 days Reduction in seedling growth, root and shoot weight
Triticum aestivum
Hydroponics Jamal et al. 2006
5, 10, 25, 50, 100 mM
14 days Reduction in plant biomass and water content, change in leaf cellular structure, decrease in chloroplasts
Brassica juncea
Hydroponics Shiyab et al. 2009
0.5, 50, 250,
500 μM 15 days
Decreased in chlorophyll content, decrease the activity of enzymes
Cucumis sativus
Hydroponics Cargnelutti et al. 2006
206
Annexure Table 4: Harmful effects of cobalt on plant growth
Co applied level
Duration Co toxic effects on plant Plant Species Growth Medium
Reference
400 µM 2 days Leaf angles decreased, decreased in hydrated osmotic potentials and water potentials
Phaseolusvulgaris
Hydroponics Rauser and Dumbroff1981
5, 50 µM 19 days Reduction in shoot weight Lycopersicon esculentum
Hydroponics Bakkaus et al. 2005
5, 50 µM 26 days Plants yellowed and yielded a smaller biomass
Triticum aestivum
Hydroponics Bakkaus et al. 2005
1,10, 50,100, 500, 1000 µM
30 days Inter-veinal chlorosis, inhibited iron transport or sequester the chelators
Nicotianatabacum
Hydroponics Nair et al. 2012
62.5, 125, 250, 500, 1000 µM
15 days Reduction in enzyme activity Lycopersicon lycopersicum
Hydroponics López-Moreno et al. 2016
20 µM 2,7 days Affected plant genes activity Arabidopsis thaliana
Hydroponics Zientara et al. 2009
5, 10, 20, 40, 50 mg kg-1
14 days Delayed in seed germination, reduction in root and shoot growth
Helianthus annuus
Hydroponics Jadia and Fulekar 2008
0.05,0. 25 mM 0, 12, 24, 48 h
Affected plant roots activity Allium cepa Hydroponics Soudek et al. 2009
50, 100, 150 µM
14 days
Decreased seed germination percentage, seedling growth, leaf area, root development and biomass production
Glycine Max Hydroponics Imtiyaz et al. 2014
0.05, 0.1, 0. 2, 0.4, 0.5 mM
3 days Chlorosis of young leaves, necrotic spots
Lycopersicon esculentum
Hydroponics Gopal et al. 2003
207
Annexure Table 5: Harmful effects of nickel on plant growth
Ni applied level
Duration Ni toxic effects on plant Plant Species Growth Medium
Reference
10, 100 µM 28 days Chlorosis, impaired growth Noccaea caerulescens
Hydroponics Visioli et al. 2014
4. 24, 21. 2,42.4 µM
30 days Chlorosis, decreased plant growth
Nicotiana Hydroponics Nair et al. 2012
10 µM 14 days Reduction in plant biomass Populus jacquemontiana
Hydroponics Mihucz et al. 2012
10µ M 32 days Leaf chlorosis and necrosis in the leaf edges
Cucumis sativus Hydroponics Csog et al. 2011
50 µM 30 days Chlorosis on new growth Alyssum murale Hydroponics Tappero et al. 2007
100 µM 4 days Reduction in leaf area and water content
Helianthus annuus Hydroponics Pena et al. 2008
25, 50, 100 µM 21 days Reduction in Chl-a and Chl-b contents
Brassica juncea Hydroponics Amari et al. 2014
25, 50, 100 µM 21 days Young leaf chlorosis, old leaf necrosis, reduction in root length and biomass
Mesembryanthemum crystallinum
Hydroponics Amari et al. 2014
10, 100, 250, 500 µM
28 days Decreased plant growth Alyssum inflatum Hydroponics Ghasemi and Ghaderian 2009
100 µM 42 days Decreased plant length, distortion in leaf stomata, increased thickness of cuticle
Lycopersicon esculentum
Hydroponics Mosa et al. 2016
3, 60, 120 µM 28 days Bleaching of the youngest leaves
Matricaria chamomilla
Hydroponics Kovacik et al. 2009
100 mg/l 28 days Decreased in leaf water potential and osmotic potential, Reduction in growth
Brassica napus Hydroponics Ali et al. 2009
0.050, 0.100, 0.175, 0.250 mM
10 days Root and shoot growth decreased
Alyssum bertolonii Hydroponics Galardi et al. 2007
20, 40, 80, 160 µM
0.5, 6, 12, 24, 48, 72, 96, 120,144 hrs
Partial necrosis, decrease in pigment contents, chlorophyll concentration and photosynthesis
Salvinia minima Hydroponics Fuentes et al. 2014
208
20 mg/l 9 hrs Decreased in root growth Alyssum bertolonii Hydroponics Nedelkoska and Doran2000
5, 30, 50, 75, 100, 130, 160, 200, 250 µM
1 hrs Effect on cell wall Leptoplax emarginata
Hydroponics Redjala et al. 2010
15,30 mg/l 30 days Reduction in chlorophyl contents and photosynthetic contents
Helianthus Annuus Hydroponics Mukhtar et al. 2010
0.04,0.6,0.7 mg/l
80 days Affected plant hormones Helianthus Annuus Hydroponics Naqvi et al. 2006
10, 25, 50, 75, 100 µM
7days Reduction in plant biomass, significant reduction in plant water loss
Thlaspi arvense Hydroponics Kramer et al. 2007
25 μmol/L 7days Decreased root and shoot fresh biomass
Arabidopsis thaliana
Hydroponics Nishida et al. 2015
50, 250 µM 0, 12, 24, 48 hrs
Decreased root growth, decreased in chlorophyll and carotenoid contents
Allium sativum Hydroponics Soudek et al. 2011
Annexure Table 6: Harmful effects of zinc on plant growth
Zn applied level
Time Duration
Zn toxic effects on plants Plant Species Growth Medium
Reference
230, 460, 1800 mg/kg
146 days Decreased plant biomass and root length
Agrostis stolonifera
Hydroponics Bernhard et al. 2005
0.01, 10 µM 21 days Biomass reduction Brassica oleracea
Hydroponics Navarro-León et al. 2016
1,2µM 60 days Reduction in plant growth, chlorophyll contents, crude proteins
Vigna radiata Hydroponics Samreen et al. 2013
10, 50, 100, 150 μmol/L
7days Chlorosis and growth reduction Solanum lycopersicum
Hydroponics Cherif et al. 2011
1, 5,10, 25, 50, 100, 125, 150 mg/l
7days Chlorosis and decreased dry matter accumulation, root death
Brassica rapa Hydroponics Coolong et al. 2004
1. 2, 50, 100, 300 μM
10 days
Metabolism shutdown, decreased in aerobic respiration and impairment of defence systems, cell death
Beta vulgaris Hydroponics Gutierrez-Carbonell et al. 2013
2.5 μM 20 days Reduction in shoot and root biomass
Salix viminalis Hydroponics Utmazian et al. 2007
209
200, 400, 800 μM
28 days Biomass reduction, chlorosis and necrosis
Pteris vittata Hydroponics Wu et al. 2009
20, 40, 80,160 mg/l
60 days Reduction in biomass and leaf size
Potentilla griffithii
Hydroponics Hu et al. 2009
0.05, 2, 6, 18, 54, 162, 456 μM
14 days Yellowing, mottling and curling Solanum tuberosum
Hydroponics Barben et al. 2007
50, 100, 300 μM
10 days
brownish root system leaf rolls, short lateral roots, decreased in carotenoid and chlorophyll contents
Beta vulgaris Hydroponics Sagardoy et al. 2009
210
SUBJECT INDEX
A
ABC-type transporter · 160
abiotic · 22
abiotic stress · 194
accumulation · 18
acidic soils · 189
active pathway · 28
Adenosine Triphosphatase · 23
adsorption · 15
Alcohol dehydrogenase · 192
alkoxyl · 46
amino acids · 203
antimitotic effect · 79
antioxidative defence system · 55
APX · 58
Arbuscular mycorrhizae fungi · 186
arsenate · 33
Arsenic · 26
Arsenic sequestration · 41
arsenite · 33
ascorbate · 55
ascorbate–glutathione · 204
ascorbate-glutathione cycle · 54
ascorbic acid · 57
ASTDR · 26
AtIRT1 transporters · 160
AtMTP3 · 203
ATPase · 23
ATP-binding cassettes · 52
ATSDR · 11
B
bioavailability · 12, 20, 23
Bioavailability · 33
biochemical · 17, 23
biofortification · 196
biogeochemical behaviour · 11
Biogeochemical behaviour · 11
biotic · 22
bulk soil · 188
C
Capermeable channels · 74
calcareous soils · 189
cancer · 25
carbohydrates · 145
Carbonic anhydrase · 192
Carboxypeptidase · 194
carcinogenesis · 25
cardiovascular disease · 25
carotenoids · 55
CAT · 55
catalytic role · 191
cation-exchange capacity · 21
CAXs · 52
CCA · 27
CDF · 51
Cellular compartmentalization · 85
Chemical speciation · 29
chlorophyll · 117
chloroplasts · 23
chlorosis · 23
chromosomal aberrations · 24
chromosomal bridges · 117
chromosomal rings · 117
chromosome fragmentation · 117
chromosome fusion · 117
chronic anaemia · 25
C-karyokinesis · 117
clay minerals · 196
Cobalt · 128
cognitive impairment · 25
Compartmentalization · 170
compartmentation · 12
complexation · 183
CoNP · 141
contaminant · 17
contaminants · 14
crops · 25
CuZn-superoxide dismutase · 193
D
desorption · 15
detoxification · 12
dicotyledonous plants · 196
dieback · 196
divalent metal transporter · 159
DMT · 109
E
edible plant · 25
EEA · 20
211
endodermis · 22
endomembrane transporters · 160
environmental · 17
enzyme inactivation · 118
enzymes · 172
EPA · 18
erosion · 18
Essential metals · 16
exchangeable · 20
exchangeable portion · 20
F
fertilizers · 21
Food safety · 25
fossil fuel burning · 21
fossil fuels · 21
G
genetic damage · 45
geological · 18
Glutathione · 53
GPX · 58
GR · 59
GSH · 170
GSSG · 170
GST · 54, 58
H
Haber-Weiss reaction · 199
Heavy metals · 18
heavy metals · 16
hemimorphite · 180
hemoglobin synthesis · 25
HMAs · 51
Hormesis · 95
hydrogen peroxide · 46
hydrolases · 191
hydroxyl · 46
hydrozincite · 183
hyperaccumulator plants · 161
I
ILZSG · 177
immobile · 73
immobilization · 22
inorganic · 22
isomerases · 191
IWGotEoCRtH · 144
J
JECFA · 112
L
labile · 183
Lead · 62
Lead toxicity · 77
ligases · 191
Light metal · 16
lime · 195
Lipoxygenase · 88
lithogenesis · 18
lyases · 191
M
malleable metal · 154
malondialdehyde · 164
MAPKs · 150
membrane permeability · 195
MerA · 108
MerB · 108
Mercury · 97
metabolic · 17
metabolic metabolism · 163
metal · 15
metalloenzymes · 162
metalloid · 16
methylation · 38
methylmercury · 107
microbial activity · 23
Microorganisms · 37
Minerals · 19
mining · 21
mobility · 20
Mobility pattern · 70
morphological · 23
MTP1 · 203
mutagenesis · 25
N
natural As levels · 31
Natural sources · 18
necrosis · 23
nervous system · 25
Ni–Cd batteries · 154
Nickel · 152
NIP · 40
nitrogen · 184
NOAEL · 95
non-essential · 28
212
non-tolerant · 197
nutrients · 137
O
organic · 22
organic acids · 203
organic hydroperoxide · 46
organic matter · 20
Organic matter · 35
oxidation states · 29
oxidative stress · 46
oxidoreductases · 191
ozon stress · 171
P
P transporter proteins · 39
passive pathway · 28
peptide cleavages · 194
peptides · 121
Peroxidase · 58
peroxyl · 46
persistent · 22
pharmaceuticals · 179
phospholipid · 194
phosphorous · 28
physico-chemical characteristics · 160
physiological · 17
phytoavailability · 22
Phytochelatin synthase · 86
Phytochelatins · 52
phytosiderophores · 184
pKa · 34
plant roots · 22
plasma membrane transporters · 160
plumule · 164
POD · 205
pollution · 11
polymetallic ores · 180
polyploid · 117
Precious metals · 16
Precipitatio · 158
primary roots · 115
protochlorophyllide oxidoreductase · 114
PSII · 114
R
radicle · 164
Reference Daily Intake · 197
remediation · 20
rhizosphere · 188
ribosomes · 194
risk · 27
RNA-polymerase · 194
ROS · 46
Rubisco · 197
RuBP-carboxylase · 193
S
Salicylic acid · 93
scavenging · 123
serpentinic · 156
S–Hg–S bridge · 115
shoots · 161
singlet oxygen · 46
smelting · 21
smithsonite · 183
soil mineralogy · 20
Soil organic matter · 35
Soil pH · 33
soil-plant · 11
solubility · 20
sorption · 40
specific gravity · 17
stomatal movement · 198
superoxide anion · 46
Superoxide dismutase · 58
T
temperature stress · 171
teratogenesis · 25
tolerant · 163
tonoplast transporters · 160
Toxic metals · 16
toxicity · 14
toxicosis · 27
Trace metals · 16
transfer · 22
transfer factors · 40
transferases · 191
Translocation · 42
transporter families · 51
tungsten carbide · 141
U
ultramafic soils · 156
uptake · 12, 22
US EPA · 26
USGS · 102
V
vacuole remobilization · 160
vacuoles · 203
vegetables · 25
213
vesicular-arbuscular mycorrhizae · 189
volcanoes · 153
W
waste incineration · 21
water stress · 171
weathering · 18
X
xylem apoplast · 162
xylem loading transporters · 160
Z
Zinc · 177
Zinc deficiency · 195
zincite · 183
ZIP · 110
ZmYS1 transporter · 161
Zn-deficiency · 196
Zn-hyperaccumulating · 203
GLOSSARY
(Source Internet)
Active pathway: It is a phenomenon by which an element passes across a membrane from an area of its low concentration to an area of high concentration by using energy. This process is generally supported by carrier proteins.
Alcohol dehydrogenases (ADH): These include a group of dehydrogenase enzymes that support the interconversion between aldehydes and alcohols or ketones with the reduction of nicotinamide adenine dinucleotide.
Alkoxyl: It is an alkyl group (carbon and hydrogen chain) singular bonded to oxygen (R–O).
Antimitotic effect: It is an effect of any substance relating to cell, which interfers the process of cell division.
Antioxidant defence system: Theis is a system present in organisms that protects organisms from toxic effects of ROS by converting them to inactive and non-toxic products.
Antioxidants: These are chemical compounds which contain scavenging groups (monohydroxy/polyhydroxy phenol); they minimize lipid peroxidation by inhibiting the oxidation of molecules.
APX: Ascorbate peroxidases are enzymes that use ascorbate as a substrate to detoxify peroxides such as hydrogen peroxide.
Arsenate: Any compound which contains AsO3− ion.
Arsenic: It is chemical element with symbol As and atomic number 33. Arsenic is a metalloid generally considered in heavy metals.
Arsenite: It is a chemical compound that contains an arsenic oxoanion with an oxidation state of +3.
Ascorbate: An ester, salt or the anion of ascorbic acid
Ascorbate-glutathione cycle: It is a metabolic pathway that detoxifies hydrogen peroxide. The cycle involves differnt antioxidant compounds such as: glutathione, ascorbate and NADPH.
Ascorbate–glutathione: The complex formed by the combination of glutathione and ascorbate during detoxification of hydrogen peroxide.
Ascorbic acid: It is a white solid naturally occurring organic compound with antioxidant properties.
ASTDR: It is the abbreviation of “Agency for Toxic Substances and Disease Registry”, which is based in Atlanta, Georgia. ATSDR is a federal public health agency of the U.S. Department of Health and Human Services.
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ATPase (adenosine triphosphatase): It is an enzyme that catalyzes the formation of ATP from ADP.
ATP-binding cassette: These are members of a transport system superfamily which are involved in the transport of heavy metals inside plants.
Bioavailability: It refres to the part of a substance which can interactwith or can be taken up by living organisms.
Biogeochemical behaviour: The behaviour or fate of an element between the physical environment and the living organisms.
Ca-permeable channels: It is is an ion channel which displays selective permeability to calcium ions.
Carbonic anhydrase: It forms a family of enzymes that catalyses the rapid interconversion of carbon dioxide and water to bicarbonate and protons, reversible reaction.
Carotenoids: It is a class of mainly orange, yellow, or red fat-soluble pigments, including carotene, which give colour to plant parts.
CAT (Catalase): It is a common enzyme found in almost all living organisms which catalyzes the decomposition of hydrogen peroxide to oxygen and water.
Cation diffusion facilitators (CDFs): These are integral membrane proteins that remove divalent metal ions from cells and thus reduce their toxicity.
Cellular compartmentalization: It is accumulation of an element in different parts of a cell.
Chemical speciation: It refers to the distribution of an element in different chemical forms in a given system
Cobalt: It is a chemical element generally considered a heavy metal with symbol Co, and atomic number 27.
Contaminant: It is an element or compound that makes a place or a substance no longer suitable for use.
DMT (N,N-Dimethyltryptamine): It is a psychedelic compound of the tryptamine family. It is a afunctional analog of psychedelic tryptamines and a structural analog of serotonin and melatonin.
European Environment Agency (EEA): It is the agency of the European Union that provides independent data/information about the environment
EPA: Environmental Protection Agency is an agency of the U.S. federal government which was established to protect environment and human health by enforcing regulations based on laws passed by Congress.
Essential metals: These are the metals which generally play an essential role in living organisms such as zinc, iron, manganese, copper, chromium etc.
Food safety: It is the protection/production of human and animal food in a way that it is is free from pollutants/contaminants.
Genetic damage: It is damage to DNA which can cause an error during DNA replication, thus causing a gene mutation that, in turn, could cause a genetic disorder.
Glutathione (GSH): It is an important antioxidant that is capable to prevent damage to important cellular components caused by ROS.
Glutathione reductase (GR): It is also known as glutathione-disulfide reductase (GSR). It catalyzes the reduction of GSSG (glutathione disulphide) to the GSH (glutathione), which is an important antioxidant in resisting oxidative stress.
Glutathione S-transferases (GSTs): It catalyzes the conjugation of the reduced form of glutathione (GSH) to xenobioticsubstrates to avoid oxidative stress.
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GPX: Glutathione peroxidase is the general name of an enzyme family, which reduces hydrogen peroxide to water.
GSSG (glutathione disulphide): It is a disulfide derived from two glutathione molecules. It is a part of Ascorbate-glutathione cycle.
Heavy metals: It is a loosely defined group of elements, which represents the elements with atomic mass generally > 20 and specific gravity > 5.
HMAs: Heavy metal ATPases (HMAs) are transporters proteins involved in the transport of heavy metals inside plants.
Hormesis: It refers to a biphasic dose response of an element characterized by a low dose beneficial effect and a high dose toxic effect.
Hydrogen peroxide (H2O2): Biologically it is considered a ROS. It palys a beneficial role of signaling molecule. Under heavy metal stress, overproduction of H2O2 induces oxidative stress.
Hydroxyl: It is a functional group containing one oxygen atom connected by a covalent bonding to one hydrogen atom.
Lead: It is a heavy metal with symbol Pb and atomic number 82.
Lipoxygenases: It is a family of iron containing enzymes most of which catalyze the dioxygenation of polyunsaturated fatty acids in lipids.
MerA (Mercury[II] reductase): It is an oxidoreductase enzyme and flavoprotein that catalyzes thereduction of Hg2+ to Hg0.
Mercury (Hg): It is a heavy silvery-white metal which is liquid at ordinary temperatures. It has an atomic number 80.
Metal: It is an element, compound, or alloy that is typically hard, shiny, opaque, and a good conductor of heat and electricity.
Metalloid: An element that has the properties of both a metal and non-metal such as arsenic, boron, silicon.
Methylation: It refers to the addition of a methyl group on a substrate.
Nickel: It is a transition metal with symbol Ni and atomic number 28.
NIP (nodulin26-like intrinsic protein): These proteins are involved in the transport of arsenite in plants.
NOAEL: No observed adverse effect level refers to the level of exposure of an organism, at which there is no biologically or statistically significant toxicity.
Organic matter: It is composed of organic compounds that have originated from the decomposition of remains of living organisms such as animals and plants.
Oxidation states: It represents the number of electrons gained or lost by an atom e.g. oxygen has -2 oxidation state.
Oxidative stress: It refers to an imbalance between the production and scavenging of free radicals, which resulst in toxicity to essential cellular molecules.
Passive pathway: It refers to the movement of an element across cell membranes without need of energy such as osmosis and diffusion.
Persistent: It refers to existence of an element over a prolonged period.
Phytochelatins: These are oligomers of glutathione, which act as chelators. These are involved in heavy metal detoxification in plants.
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Reactive oxygen species (ROS): These are chemically very reactive oxygen species containing an extraelectron. These species can induce oxidative stress when their production increases to toxic level.
Soil pH: It refers to the negative of H+ ion activity. It is a measurement of the acidity or alkalinity of a soil. On the pH scale, 7.0 is neutral.
Sorption: It is a physical and chemical process by which one substance becomes attached to another.
Superoxide dismutase (SOD): It is an enzyme that catalyzes the convertion of superoxide (O2−)
radical into hydrogen peroxide or oxygen.
Translocation: It refers to the movement of metal ions within plants (roots to shoot or shoot to roots).
USGS: The United States Geological Survey is a scientific agency of the United States government, which provides data regarding natural resources, and the natural hazards.
Zinc: Zinc is a chemical element generally considered heavy metal with symbol Zn and atomic number 30.
ZIP: It is a metal transporter family which is capable of transporting a variety of cations in plants.
