Metals in the Environment: Analysis by Biodiversity (Books in Soils, Plants, and the Environment)

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Preface

Metal contamination and pollution in the environment, the significance of biodiversity conservation, and the root causes of biodiversity loss are emphasized in a wide array of works. However, information about the utility of phytodiversity and microbial diversity for the management of metal-contaminated and -polluted ecosystems is rather scanty. Therefore, the collective work presented in this volume spotlights the potentiality of biodiversity for monitoring and abatement of metal pollution in the environment, and also explores the emerging issues and initiatives concerning metals in the environment. Heavy metals are being enriched in all aspects of the environment, viz., air, water, and soil, by anthropogenic as well as natural processes. For example, effluents from industrial and mining wastes have increased several thousand-fold the metal concentrations in river water and sediments. Aquatic systems (freshwater, marine, and estuarine) act as receptacles for several metals. Serpentine soils contain heavy metals for nickel (averaging 10 mg per g of soil), as well as cobalt and chromium—both of which are present at lower levels than nickel and are characterized by high concentrations of iron and magnesium and low nutrient levels. Some of the best-studied nickeliferous biotopes are the New Caledonian soils, which are inhabited by nickel-resistant bacteria as well as metal-tolerant plants, many of which act as nickel hyperaccumulators. An interesting ecosystem is established in these biotopes driven by a nickel cycle, in which hyperaccumulating trees extract nickel from deep soil and rock layers and subsequently store it in their leaves (up to 1% Ni in leaf dry matter). When the leaves are shed from the trees, the nickel is leached out into the surrounding topsoil. The solubilized metal exerts a localized selective pressure on the topsoil microflora, which aciii

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quire resistance to high levels of nickel (⬎20 mM). In fact, the predominant cultivable microflora found in the topsoil (bacteria, fungi, and protozoa) adapt to the high levels of nickel. Interestingly, the microflora that were not found directly beneath the canopy but in the same soil showed tolerance to lower levels of nickel (3 mM) compared to the resistant population. Thus, the nickel selection pressure existing as a gradient around the hyperaccumulator plants has a dramatic effect on the composition of the local microbial population. The most successful monitoring methods for metals in the environment are based on gene- and proteinbased bacterial heavy metal biosensors. Likewise, several associations of mosses, liverworts, and ferns are capable of growing on metal-enriched substrates. Each of these groups of plants possesses certain anatomical and physiological properties that enable representatives to occupy unique ecological niches in natural and man-made metalliferous environments. The best documented of these are the groups of specialized bryophytes found on substrates enriched with copper; so-called ‘‘copper mosses’’ are found worldwide and come from widely separated taxonomic groups. Other bryophytes are associated with lead- and zinc-enriched substrates. No distinctive bryophyte flora were observed on serpentine soils. Pteridophytes are associated with serpentine substrates in various parts of the world, and several African species grow on copper-enriched soils or on soils polluted by metal-smelter emissions. The existence of these diverse groups is a clear indication of the ability of bryophytes and pteridophytes to adapt to extremes of metal content of their growth substrate, either by avoidance of the toxic constituents or by expressing resistance/tolerance to metals at an organismal or cellular level. Among angiosperms, the hyperaccumulators are reported in Brassicaceae, Cyperaceae, Cunoniaceae, Caryophyllaceae, Fabaceae, Flacourtiaceae, Lamiaceae, Poaceae, and Violaceae. About 400 metal-hyperaccumulating species have been identified, and this number might increase in the near future. The consequences and ecological role of metal hyperaccumulation are expanding rapidly. Different schools of thought have their own convictions. The salient views are: Hyperaccumulation is possible when plants have hypertolerance properties that may be the result of chelation and vacuolar compartmentalization. In general, the root of a plant contains more metal than shoot. In plants that hyperaccumulate heavy metals, the roots must transfer the bioaccumulated metals to the shoots. It has been reported that in normal plant root, Zn, Cd, or Ni concentrations are 10 or more times higher than shoot concentrations, but in hyperaccumulators, shoot metal concentrations may exceed those in roots. Mn and Zn are reported to concentrate up to 1% and more in leaf dry matter in the case of hyperaccumulators.

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Hyperaccumulators can resist drought and can interfere with neighboring plants. It is also suggested that hyperaccumulators provide protection against fungal and insect attack. Recent studies suggests that Ni hyperaccumulation has a protective function against fungal and bacterial pathogens in Streptanthus polygaloides and folivores in S. polygaloides and Thlaspi montanum. An antiherbivory effect of Zn has been found in the Zn hyperaccumulator T. caerulescens. The fundamental aspects of microbe/plant stress responses to different doses of metals coupled with breakthrough research innovations are amenable for advancing environmental biotechnology. Accordingly, the chapters in this work are organized into two parts: Chapters 1–9 are based on taxonomic/habitat criteria: bacteria, mycorrhizal fungi, freshwater algae, saltmarsh metallophytes, lichens, bryophytes and pteridophytes, angiosperms (Brassicaceae, Caryophyllaceae, Fabaceae, and Poaceae [the tolerance of monocots to heavy metals is generally higher than that of dicots and requires a rather long exposure time to heavy metals]), and aquatic macrophytes. Chapters 10–15 are based on functionality and utility of biodiversity: angiospermous and gymnospermous tree rings and dendroanalysis, tree crops, tree bark, heavy-metal interactions with soil microbes, consequent implications, subsoil acidification and plant responses to aluminum, and heavy-metal behavior in soils emphasizing certain remediation strategies. Tree rings are the result of a rhythmic activity of the cambium. In spring and early summer the cambium is active and new wood is formed, while in autumn and winter the cambium is dormant. The mineral composition of the tree rings depends partly on the mineral uptake of the tree. The extent of mineral uptake may reflect the availability of the elements in the tree’s environment. In fact, several of the gymnosperms (Cedrus, Chamaecyparis, Cryptomeria, Gingko, Juniperus, Larix, Pinus, Pseudotsuga, Picea, Taxodium, Thuja, and Tsuga) are being used as retrospective biomonitors of trace-metal pollution. Bark enclosed in tree trunks, known generally as tree ‘‘bark pockets,’’ provides invaluable information for monitoring metal pollution in the atmosphere. Because bark pockets were not previously considered useful parts of a tree, tree trunks with bark pockets were discarded as waste. Many forest trees with bark pockets have therefore been left standing and not harvested. Several plants and microbes are being employed for remediation of the metal-polluted ecosystems. For example, constructed wetlands, reed beds, and

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floating-plant systems are quite common for the treatment of wastewaters and industrial effluents containing toxic heavy metals. In situ and on-site remediation of metallic residues is gaining considerable global significance. As civilization progresses we need to conserve biodiversity for biomonitoring and decontaminating the environment. I fervently believe that this work will stimulate future research in the enchanting field of metals in the environment, leading to progress in conserving the unknown for the prosperity of humankind. I am extremely grateful to Padma Bhushan Professor P. Rama Rao, ViceChancellor, University of Hyderabad, for inspiring me to focus my research on biodiversity and for his constant encouragement. I thank Professor T. Suryanarayana, Dean, School of Life Sciences, for supporting my academic endeavors. I am grateful to all the contributors for comprehensive reviews which culminated in the present text. I am pleased with the excellent cooperation of my wife, Savithri. The superb and skillful job of Ms. Annie Cok and timely action from Mr. Jeff Stockton, Ms. Marilyn Ludzki, and Mr Russell Dekker at Marcel Dekker, Inc., produced this work in a timely manner. M. N. V. Prasad

Contents

Preface

iii

Contributors

ix

1

Bacteria Daniel van der Lelie and C. Tibazarwa

1

2

Mycorrhizal Fungi Jan V. Colpaert and P. Vandenkoornhuyse

37

3

Freshwater Algae Barbara Pawlik-Skowron´ska and Tadeusz Skowron´ski

59

4

Salt Marshes Isabel Cac¸ador and Carlos Vale

95

5

Lichens Cristina Branquinho

117

6

Bryophytes and Pteridophtes Nicholas W. Lepp

159

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viii

7

Contents

Angiosperms (Asteraceae, Convolvulaceae, Fabaceae and Poaceae; other than Brassicaceae) Anna Siedlecka, Anna Tukendorf, Ewa Sko´rzyn´ska-Polit, Waldemar Maksymiec, Małgorzata Wo´jcik, Tadeusz Baszyn´ski, and Zbigniew Krupa

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8

Brassicaceae Luigi Sanita` di Toppi, Maria Augusta Favali, Roberto Gabbrielli, and Patrizia Gremigni

219

9

Aquatic Macrophytes M. N. V. Prasad, Maria Greger, and Bruce N. Smith

259

10

Aluminum Toxicity in Acid Soils: Plant Responses to Aluminum Hideaki Matsumoto, Yoko Yamamoto, and S. Rama Devi

11

Tree Crops Tracy Punshon

12

Tree Bark: Tree ‘‘Bark Pockets’’ as Pollution Time Capsules for Historical Monitoring Kenichi Satake

13

Tree Rings and Dendroanalysis C. Nabais, J. Hagemeyer, and H. Freitas

14

Heavy Metal Interactions in Soils and Implications for Soil Microbial Biodiversity R. Naidu, G. S. R. Krishnamurti, N. S. Bolan, W. Wenzel, and M. Megharaj

15

Behavior of Heavy Metals and Their Remediation in Metalliferous Soils Arun B. Mukherjee

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353

367

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Biodiversity Index

473

Subject Index

483

Contributors

Tadeusz Baszyn´ski Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland N. S. Bolan Institute of Natural Resources, Massey University, Palmerston North, New Zealand Cristina Branquinho Oeiras, Portugal

Environmental Department, Universidade Atlaˆntica,

Isabel Cac¸ador Institute of Oceanography, University of Lisbon, Lisbon, Portugal Jan V. Colpaert Center for Environmental Sciences, Limburgs Universitair Centrum, Diepenbeek, Belgium S. Rama Devi Research Institute for Bioresources, Okayama University, Kurashiki, Japan Maria Augusta Favali Department of Evolutionary and Functional Biology, Section of Plant Biology, University of Parma, Parma, Italy H. Freitas Departamento de Botaˆnica, Universidade de Coimbra, Coimbra, Portugal ix

x

Contributors

Roberto Gabbrielli Department of Plant Biology, University of Florence, Florence, Italy Maria Greger Department of Botany, Stockholm University, Stockholm, Sweden Patrizia Gremigni Center for Legumes in Mediterranean Agriculture (CLIMA), University of Western Australia, Perth, Western Australia, Australia J. Hagemeyer Department of Ecology, University of Bielefeld, Bielefeld, Germany G. S. R. Krishnamurti Remediation of Contaminated Environments Program, CSIRO Land and Water, Urrbrae, South Australia, Australia Zbigniew Krupa Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Nicholas W. Lepp School of Biological and Earth Sciences, Liverpool John Moores University, Liverpool, United Kingdom Waldemar Maksymiec Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Hideaki Matsumoto Kurashiki, Japan

Research Institute for Bioresources, Okayama University,

M. Megharaj Remediation of Contaminated Environments Program, CSIRO Land and Water, Urrbrae, South Australia, Australia Arun B. Mukherjee Department of Environmental Protection, University of Helsinki, Helsinki, Finland C. Nabais Departamento de Botaˆnica, Universidade de Coimbra, Coimbra, Portugal R. Naidu Remediation of Contaminated Environments Program, CSIRO Land and Water, Urrbrae, South Australia, Australia Barbara Pawlik-Skowron´ska Sciences, Lublin, Poland

Institute of Ecology of the Polish Academy of

Contributors

xi

M. N. V. Prasad Department of Plant Sciences, University of Hyderabad, Hyderabad, India Tracy Punshon Advanced Analytical Center for Environmental Services, Savannah River Ecology Laboratory, Aiken, South Carolina Luigi Sanita` di Toppi Department of Evolutionary and Functional Biology Section of Plant Biology, University of Parma, Parma, Italy Kenichi Satake Department of Global Environment, National Institute for Environmental Studies, Tsukuba, Japan Anna Siedlecka Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Ewa Sko´rzyn´ska-Polit Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Tadeusz Skowron´ski Institute of Ecology of the Polish Academy of Sciences, Lublin, Poland Bruce N. Smith Department of Botany and Range Science, Brigham Young University, Provo, Utah C. Tibazarwa Department of Environmental Technology, Flemish Institute for Technological Research (Vito), Mol, Belgium Anna Tukendorf Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Carlos Vale Research Institute for Fisheries and Sea Research, Lisbon, Portugal P. Vandenkoornhuyse Laboratorie de Ge´ne´tique et Evolution des Populations ve´ge´tales, Universite´ de Lille, Paris, France Daniel van der Lelie Department of Environmental Technology, Flemish Institute for Technological Research (Vito) Mol, Belgium W. Wenzel Remediation of Contaminated Environments Program, CSIRO Land and Water, Urrbrae, South Australia, Australia and University of Agricultural Sciences Vienna (BOKU), Vienna, Austria

xii

Contributors

Małgorzata Wo´jcik Department of Plant Physiology, Maria Curie-Skłodowska University, Lublin, Poland Yoko Yamamoto Research Institute for Bioresources, Okayama University, Kurashiki, Japan

1 Bacteria Daniel van der Lelie and C. Tibazarwa Flemish Institute for Technological Research (Vito), Mol, Belgium

1

INTRODUCTION

This chapter focuses on the way in which bacteria respond to heavy metals in the environment, either from natural sources or due to anthropogenic activities. Metal resistance in bacteria is an area whose potential exploitation in the areas of bioremediation, environmental monitoring, and environmental cleanup technologies is progressively being realized. The inclusion of microorganisms and their products in metal recovery and environmental biotechnologies has been the subject of ever-increasing attention during the past few decades. Biologically, bacterial heavy metal resistance mechanisms and their regulation are interesting phenomena in themselves, since many metals cations are necessary requirements in cellular metabolism; thus, in the presence of abnormally high levels of metal cations, the bacteria must strike a fine balance between uptake of the required amounts of metal cations and expulsion of the excess. This is referred to as heavy metal homeostasis.

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van der Lelie and Tibazarwa

ECOLOGICAL RELEVANCE OF BACTERIAL INTERACTIONS WITH HEAVY METALS

Bacteria have been interacting with heavy metals since their early evolutionary history. Bacterial interactions with heavy metals have been best studies in extreme environments where the emphasis has been to assess bacterial adaptation, metabolism, tolerance, and resistance to heavy metals. In general, the types of interactions are directly dependent on the biological role of the heavy metal species in the bacteria: some heavy metals are required as essential cofactors for protein activity or to stabilize protein conformations, while most heavy metals tend to be toxic at high concentrations (reviewed in 1). Under these extreme environments, a selective advantage is conferred to those organisms that have adopted resistance mechanisms to withstand the toxic effects of high concentrations of heavy metals. For a given biotope or ecological niche, these types of interactions are also indirectly affected by the utilization of heavy metals by other organisms within that biotope. Within the dynamics of ecological evolution, the biological availability (bioavailability) and speciation of heavy metals can act as the main selective pressure determining the composition of the microbial population. It has been observed for a number of biotopes that if one or more commensals of the biotope are able to immobilize the heavy metal species in that environment, then the biological inactivation of the metal species would reduce its bioavailability, thereby relieving the metal selection pressure. As a result, newer or alternative populations that would not necessarily require resistance traits to survive could then flourish on this biotope. Thus, the speciation of heavy metals in a biotope determines the nature of their interaction with bacteria, which, in turn, may be wholly influenced by their interaction with other organisms. This phenomenon has been described for biotopes of both natural and anthropogenic (man-made) origin. 2.1

Natural Metal-Rich Biotopes

Examples of natural metal-rich biotopes include serpentine (or ultramafic) soils, which commonly result from the weathering of serpentinite rocks (2). These soils tend to be enriched in the heavy metals nickel (averaging 10 mg per gram soil), cobalt, and chromium—the last two of which are present at lower levels than the first—and are typified by high concentrations of iron and magnesium and low nutrient levels (2,3). Some of the best studied nickeliferous biotopes are the New Caledonian soils (4), which are inhabited by nickel-resistant bacteria as well as metal-tolerant plants, many of which act as nickel hyperaccumulators (3). An interesting ecosystem is established in these biotopes driven by a nickel cycle, in which hyperaccumulating trees extract nickel from deep soil and rock layers and subsequently store it in their leaves (up to 1% Ni in leaf dry matter). When

Bacteria

3

the leaves fall off the trees, the nickel is leached out of the plant tissues into the surrounding top soil; the solubilized metal exerts a localized selective pressure for a top soil microflora resistant to high levels of nickel (⬎20mM). In fact, the predominant cultivable microflora found in the top soil—from bacteria to fungi and protozoa—are resistant to high levels of nickel (4). Interestingly, the microflora, which was not found directly beneath the leaf canopy but in the same soil, showed tolerance to lower levels of nickel (3 mM) in comparison with the resistant population. Thus, the nickel selection pressure that exists as a gradient around the hyperaccumulator plants has a dramatic effect on the composition of the local microbial population. 2.2

Biotopes Rich in Heavy Metals Resulting from Anthropogenic Activities

A different situation exists in metal-rich anthropogenic biotopes, resulting from man’s activities, since the metal selective pressure is not necessarily constant. Such biotopes are found to occur in soils close to leaching activities for mining of metal ores (5) and also in the vicinities of metal-processing factories (6–8). The anthropogenic sites tend to contain particularly high metal concentrations in the range of 103 –105 mg/kg soil (see also 7,9). Some of the best studied are the pyrite-like biotopes, in which the interaction of Thiobacillus bacteria with iron and other heavy metals is directly related to a sulfur cycle, necessary for the maintenance of a balanced ecosystem within this biotope. The predominant and best characterized species in this genus is Thiobacillus ferrooxidans, which is widely used in biological mining operations (reviewed in 5,10,11). This autotrophic obligate chemolithotroph derives its energy and reducing power from the oxidation of ferrous iron and reduced sulfur compounds under strict acidic (low pH) conditions (11). All thiobacilli are capable of oxidizing reduced sulfur compounds to meet their energy requirements: T. ferrooxidans utilizes ferrous iron as an electron donor, and the key enzymes require sulfate ions as cofactor components. This organism is a commensal of pyrite-like (iron-rich) biotopes where it plays an important role in the cycling of Fe and S (11). This cycling involves the oxidation of reduced sulfur compounds—including elemental sulfur, sulfides, and sulfites—to sulfuric acid. Pyrite-like biotopes, which contain high levels of insoluble iron sulfides, can be oxidized by the bacteria to yield soluble iron sulfates, thereby ‘‘leaching’’ the metal into solution (12). This process also occurs in the presence of other minerals, including nickel, copper, lead, gallium, zinc, cobalt, and copper, and can result in the release of high concentrations of soluble heavy metal fractions into the surrounding environment. In the case of iron, the organism deals with the change in the biological availability of this cation by generating and regenerating the oxidant Fe3⫹ from Fe2⫹, which can be utilized for its own energy production, and as a micronutrient by other commensals of

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this biotope. The ‘‘leaching’’ metabolism of this organism can be detrimental to the organism if the solubilized metal cations reach toxic concentrations or if extremely toxic cations, such as arsenic present in arsenopyrite minerals, are released in a more bioavailable form. To overcome the environmental constraints caused by these heavy metals, Thiobacillus spp. have evolved resistance mechanisms to a variety of toxic metals, including arsenic and mercury. For example, the resistance to mercury volatizes the metal thereby removing it from solution. Nevertheless, it is clear that although the organism has a significant ecological advantage in being able to solubilize and consequently make trace elements available for its metabolism, this same property can have detrimental effects on its own proliferation and those of other commensals in the ecosystem if a foreign selective pressure was to be introduced into the ecosystem, such as that posed by increased contamination of a toxic heavy metal. The sulfate-reducing bacteria are found in different biospheres of both natural origin, such as thermal sulfur springs, ocean and sea beds), and of anthropogenic origin, such as mining waters from metal sulfide deposits (reviewed in 13). This group includes Desulfovibrio spp., Desulfobulbus spp. (14), Desulfobacteriaceae and other strains within the δ-proteobacteria, gram-positive species as well as archaebacteria, and are also implicated in the sulfur cycle (their phylogeny is presented in 15 and 16). The Desulfovibrio bacteria are chemoorganotrophic, strictly anaerobic, gram-negative organisms that are able to utilize sulfates and other partially oxidized forms of sulfur for anaerobic respiration. The resulting metabolite is hydrogen sulfide, which reacts easily with heavy metal cations forming metal sulfides of low solubility—the inverse metabolism to that of the thiobacilli. In addition, various sulfate-reducing bacteria can directly reduce metal oxianions, such as arsenate and chromate, that are subsequently precipitated as metal-sulfate complexes. Thus, these two groups of organisms have opposite effects on heavy metal solubility and therefore biological availability: the thiobacilli and other sulfur-oxidizing bacteria increase the bioavailability by leaching out heavy metal sulfates whereas the sulfate-reducing bacteria reduce the bioavailability by forming poorly soluble metal sulfides, which can mineralize out of solution with organic matter. In acid mine drain waters, the occurrence of both types of bacteria would affect the speciation of heavy metals, which in turn is dictated largely by the Fe and S element cycles, and would therefore influence the biological makeup of the ecosystem. 2.3

Sites of Metallurgical Industries

Bacterial interactions with heavy metals in contaminated sites resulting from metallurgical processing and related industries has been a major area of interest during the past decades, particularly in industrialized nations where the unclean practices of the metallurgy industry during the 19th and 20th centuries resulted in a

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legacy of large areas of soil contamination that are unsuitable for life. The Maatheide soil in Lommel, Belgium is one such example, characterized by heavy contamination with zinc (7000–40,000 ppm), aluminum (39,000 ppm), lead (2000 ppm), cadmium (20 ppm), and copper (1000 ppm) (6,8). In these soils, Zn was found to be highly abundant. It was also the most biologically available heavy metal and the main selection pressure for bacteria inhabiting this biotope (17). The microbial diversity of these soils has been monitored periodically during the past 12 years. Early studies revealed that the majority of bacteria on this biotope were metal-resistant Ralstonia eutropha-like strains (40% of the population), showing high levels of zinc resistance. Samples taken 10 years later revealed that this subgroup made up only 1–4% of the population and had been replaced by other gram-negative and even gram-positive bacteria that showed reduced resistance to heavy metals, including zinc (8). The progressive decrease in zinc toxicity of these soils as a result of natural weathering processes, man’s attempts at remediation, as well as the microbial interaction with the heavy metals has been attributed to this succession in microbial populations. The Ralstonia eutropha-like strains, of which the type strain CH34 is the best characterized, are gram-negative (β-protobacteria) soil bacteria displaying resistance characteristics to a variety of heavy metals, and are thus well suited for these types of biotopes (18–20, reviewed in 21–23). As described later in this chapter, the interactions of these bacteria with heavy metals can result in alteration of the metal speciation as well as its bioavailability to other organisms. Similar types of heavy metal resistance, such as the czc cobalt, zinc, and cadmium–resistant determinant, are found in Ralstonia eutropha DS185, the dominant heavy metal–resistant Ralstonia eutropha strain isolated from the Maatheide (8,20). The reduced bioavailability of Zn, Cd, Ni, Co, and Pb has relieved the selection pressure on the microbial populations so that, with time, what were the pioneering and dominant metalloresistant bacteria have been succeeded with metallotolerant species (8,20). Eventually, these metallotolerant populations may be succeeded by metallosensitive organisms, which are already present in the soil, albeit in low numbers. This was demonstrated during in situ heavy metal inactivation experiments using the soil additives beringite and steelshots to immobilize the bioavailable fraction of heavy metals present in Maatheide (17). For untreated soils, a strong increase in the metal-resistant subpopulations was observed with increasing metal contamination. However, 12 months after the application of the soil additives, up to a 100-fold decrease in the heavy metal–resistant subpopulations was observed for beringite- and beringite-plus steelshots-treated soils. However, treatment with other soil additives resulted in hardly any decrease of the heavy metal–resistant subpopulations. It was hypothesized that the observed decrease in heavy metal– resistant bacterial subpopulations was directly reflecting the efficacy of in situ metal immobilization. Further physicochemical and biological evaluation confirmed the sustainability of zinc and cadmium immobilization by beringite and

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beringite plus steelshot treatment, while the inefficient heavy metal immobilization by the other soil additives was also confirmed. This shows that the ecology of heavy metal–resistant soil bacteria, which is directly affected by the bioavailability of the heavy metals, can be used as a tool to predict the efficacy of in situ heavy metal immobilization. 3

HEAVY METAL–RESISTANT BACTERIA

The speciation of a heavy metal determines the resultant interaction with bacteria. Being able to withstand high concentrations of bioavailable heavy metals clearly offers an advantage for surviving harsh environmental conditions caused by increased heavy metal concentrations, and it is not surprising that metal resistance determinants and metal tolerance capabilities are widespread in different bacterial genera. Strains isolated from natural and anthropogenic biotopes have been assigned to the gram-negative β-proteobacteria, which include the Ralstonia eutropha strains (now termed Ralstonia metallidurans, and formerly Alcaligenes eutrophus—see 20), and γ-proteobacteria including many Pseudomonas strains. Both the β- and γ-proteobacteria exhibit heavy metal resistance phenotypes varying widely in the micromolar to millimolar range, which can be of plasmid or chromosomal origin. However, the highest resistance levels are found in members of the β subgroup (reviewed in 7, 21). The other classes isolated from anthropogenic biotopes include low-GC gram-positive species, which show heavy metal resistance in the micromolar concentration range, and some high-GC content gram-positive species (e.g., Arthrobacter spp.; 24), which can exhibit resistance phenotypes in the low millimolar range. The heterogeneous group termed Burkholderia (which are still awaiting proper classification using a molecular taxonomy approach) predominates in some natural nickeliferous biotopes, although other strains have also been isolated, belonging to the genus Acinetobacter, Agrobacterium, Comamonas, and Arthrobacter (24). However, the highest heavy metal resistance phenotypes are found in the archaebacteria and in the Thiobacillus genus, which are mostly acidophilic oligate chemolithotrophs and which have been isolated from highly polluted sites where metal concentrations can reach molar levels and are highly bioavailable due to the low pH (⬃2) conditions. Heavy metal resistance has now been characterized in many gram-negative and gram-positive bacteria, particularly those with a high-GC content, and, to a lesser extent, in the archaebacteria. Microbiological data on the types and numbers of bacterial populations in an environmental sample has been used with some degree of accuracy to assess the heavy metal pollution of the sample (reviewed in 7,17). However, due to the inaccuracies of sampling from the environment as well as the complex nature of the endogenous microbial populations, strategies are being pursued that rely principally on understanding the bacteria-heavy metal interactions at the molecular level. The interactions between bacteria and heavy metals have been widely ex-

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ploited: the most successful applications include the utilization of the acidophilic, sulfur-oxidizing T. ferrooxidans in leaching activities for the extraction of various heavy metals from low-grade ores (reviewed in 5,10). The resistance mechanisms of some other bacteria, particularly the β- and γ-proteobacteriaceae, have also been used as the underlying principles for the development of various environmental technologies: therefore, these will be discussed in greater detail. 3.1

Bacterial Heavy Metal Resistance Mechanisms

The physiological role of the metal is a key determining factor in its fate when it is encountered by a bacterium. Some metals are tolerated at high levels because they are (bio)chemically unreactive or simply do not pose a toxicity threat even at high concentrations, whereas the majority exhibit some sort of toxicity above threshold concentrations (reviewed 1,25–27). Metal toxicity may result from the fact that (1) high concentrations of some cations can competitively inhibit the normal functioning of analogous metals that are essential for cellular metabolism, e.g., as enzyme cofactors; (2) some metals are not tolerated even at low concentrations because they can interact with cellular components and damage them; (3) they can form strong interactions with key proteins in the cell, thereby inhibiting or hyperstimulating their activity; (4) DNA-metal complexes can damage the structural integrity of DNA, which can induce mutations. Bacteria have adopted the following mechanisms to deal with potentially toxic concentrations and types of metal species (also reviewed in 28): 1. Pumping out the metal before it can accumulate in the bacterial cytoplasm, typically involving an elaborate protein pump system in the membrane. 2. Actively taking up the metal by specific uptake proteins that immediately channel the toxic metal to a specific detoxification pathway. 3. Exclusion of the metal by a permeability barrier; exclusion can be total or partial. 4. Enzymatic conversion of the metal species to a less toxic form (detoxification) or into an alternative form that can activate rapid removal from the cell. 5. Sequestration or immobilization of the metal in a compartment of the cell or on the outer cell. Examples of how these mechanisms can interact to confer resistance to specific heavy metal cations in bacteria are summarized in Table 1 and will be described below. It is important to note that many of these mechanisms are encoded by extrachromosomal elements, mainly plasmids or on mobile genetic elements, and these are generally the best characterized. Nevertheless, chromosomally encoded resistance mechanisms have also been described in the literature. Some organisms have evolved both chromosomal and plasmid-encoded resistance mechanisms

TABLE 1 Examples of Bacterial Resistance Mechanisms to Heavy Metalsa

Ni

Zn

Cd

Co

Hg

Pb Cr

As (Sb) a

ATPases

CDF

Enzymatic reduction

Sequestration

Undefined efflux

Others

cnrYXHCBA nccYXHCBAN nreAB czcNICBADRSE czrRSCBA zntRA ziaRA cadCA smtBA czcNICBADRSE czrRSCBA cadCA smtBA nccYXHCBA cnrYXHCBA czcNICBADRSE czrRSCBA coaRA silPCBARSE copSRABCDGFH copABCDRS pcoABCDRS copYZAB merRTPCAD merRTPAD merRTPABD pbrTRABCD cadCA chrDCAB chrAB chrAB arsRDABC arsRBC

CnrCBA NccCBA X CzcCBA CzrCBA X X X X CzcCBA CzrCBA X X NccCBA CnrCBA CzcCBA CzrCBA X SilCBA X X X X X X X X X X X X X X

X X X X X ZntA ZiaA CadA X X X CadA X X X X X CoaA SilP CopF X X CopA, CopB X X X PbrA CadA ChrD X X ArsA X

X X X CzcD X X X X X CzcD X X X X X CzcD X X X X X X X X X X X X X X X X X

X X X X X X X X X X X X X X X X X X X X X X X MerA MerA MerA X X X ? ? ArsC ArsC

? ? X CzcE X X X X SmtA CzcE X X SmtA X ? CzcE X X SilE CopH X PcoE X X X X PbrD X X X X X X

X X NreAB X X X X X X X X X X X X X X X X X X X X X X X X X ChrA ChrA ChrA ArsB ArsB

X X X X X X X X X X X X X X X CzcN, I X X X CopABCD, G CopABCD PcoABCD X MerT, P, C MerT, P MerT, P, B PbrT, B, C X ChrB, C, D ChrB ChrB X X

Genes present in the operon but not specified in the table are involved in regulation.

Genomic localization Plasmid Plasmid Plas./chrom. Plasmid Chromosome Chromosome Chromosome Plasmid Chromosome Plasmid Chromosome Plasmid Plasmid Plasmid Plasmid Chromosome Chromosome Chromosome Plasmid Plasmid Plas./chrom. Chromosome Plas./chrom. Plas./chrom. Plas./chrom. Plasmid Plasmid Plasmid Plasmid Chromosome Plasmid Plas./chrom.

Ref. 21,35 21,36 36,67,68 32,62 37 56,104 58 25,27 94 32,62 37 25,27 94 21,36 21,35 32,62 37 105 38 22 60,61 60,61 59,60 70 70 70 55 25,27 62 84 83,85 27,75 27,109

van der Lelie and Tibazarwa

Ag Cu

Proton/ cation antiporter

8

Cation

Operon structure

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even for the same metal; however, the physiological interplay between both types of systems is rarely well understood, and in this chapter chromosomally and plasmid-encoded systems will be considered as separate, unless identical genes are involved. The evolutionary significance of chromosomal, plasmid, mobile element localization of resistance operons is assessed in 28. Ralstonia eutropha CH34 is perhaps a model organism in terms of bacterial resistance mechanisms and interaction with heavy metals in soils. CH34 is the type strain of a group formerly known as Alcaligenes eutrophus but whose taxonomic classification was recently revised to Ralstonia eutropha (20) and is currently being renamed to Ralstonia metallidurans (Mergeay, personal communication). This strain and its derivatives were originally isolated from the metal-rich sediments of a zinc factory in Belgium (18). One of the most remarkable properties of CH34 is its ability to thrive in metal-rich environments: this subject has been extensively reviewed, most recently in (9,21,22). In order to survive in metal-rich biotopes, CH34 has evolved an astounding number of genetic determinants that encode for resistance to many heavy metals and metal compounds (18,21,22). These determinants, which are mainly localized on two megaplasmids, pMOL28 (18,29) and pMOL30 (18), include resistance determinants to combinations of cobalt and nickel (cnr), cadmium, zinc, and cobalt (czc), chromate (chr), mercury (mer), thallium (tll), copper (cop), lead (pbr), and others awaiting definition. The resistance mechanisms of CH34 and other metal-resistant bacteria, as well as their genetic organization and regulatory control, are detailed below. The interactions with essential metals, including Co, Cu, Ni, and Zn, are differentiated in comparison with interactions with the highly toxic metals and metalloids such as Ag, Cd, Cr, Hg, Pb, Sb, and Bi. 3.1.1

Three-Component Cation/Proton Antiporter Systems for Metal Efflux

These systems are members of the major facilitator superfamily of transporters (reviewed in 69). The model system for three-component cation/proton antiporters is the Czc system of R. eutropha CH34, which confers resistance to cadmium, zinc, and cobalt by pumping out these metals when their intracellular concentrations exceed a certain threshold (22,30–33). This ‘‘metal pump’’ is encoded by the structural gene czcCBA. Other examples of three-component cation/proton antiporters involved in heavy metal resistance include CnrCBA, mediating efflux of cobalt and nickel in CH34 (34,35), NccCBA, with substrates nickel, cobalt, and cadmium in R. eutropha strain 31A (36), CzrCBA, with substrates zinc and cadmium in P. aeruginosa (37), and SilCBA mediating efflux of silver in Solmonella typhimurium (38). The mechanism of efflux has been described for the Czc system (22,30– 32,39,40), and it has been demonstrated to be essentially a zinc efflux system.

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Three membrane proteins are involved in transporting the divalent cations— zinc, cadmium, cobalt—across the inner membrane. The cytoplasmic membranespanning ‘‘A’’ protein, which constitutes the actual cation/proton antiporter, is a member of the RND (resistance/nodulation/division) superfamily of transporter proteins (39–43). Typical of the RND proteins, the A protein translocates the metal cation from the cytoplasm in one channel with the concomitant uptake of protons from the cytoplasm into the periplasm through a separate, adjacent ‘‘tunnel’’ (27,31,40,41). This system of transport of a substrate differs from the ATPdriven mechanisms of ATP-binding cassettes (ABC transporters; 44) where the driving force for metal cation transport comes from the translocation coupling of uptake of protons and expulsion of cations into the periplasm. The CzcA/ CzrA/NccA/CnrA proteins are highly homologous, showing conservation in the regions corresponding to the proton tunnels; however, the metal translocating region has evolved specifically for the cation substrate. The proton uptake results in an increased pH in the extracellular environment, which is related to the metal concentration, and for R. eutropha CH34 alkanization of the extracellular medium contributes to the extracellular immobilization of the metal cations. The second protein in the cation translocation pathway is the periplasmic ‘‘B’’ protein (which is anchored at its N terminus to the outer membrane), which has the specific function of ensuring the outward passage of the metal cations from the cytoplasm across both membranes, without leakage into the periplasm (31,41). The B protein is a member of a family known as membrane fusion proteins, or MFPs (31,41,45) and is functional as a dimer in the CzcCB2A complex (31). Therefore it is hypothesized that all B subunits of related metal resistance exporters would also be functional as dimers. The final passage of the metal cations through the bacterial membrane is mediated by the ‘‘C’’ protein. The latter is a member of the outer membrane factors or OMF family (41,43,45), which is anchored to the outer membrane and protrudes into the periplasm (31). The function of this protein is to translocate the metal cation from the MFP B protein toward the cell exterior. In Czc-mediated efflux, an outer membrane protein (Omp) protein is believed to carry out the terminal step in removing the cation from the outer membrane surface (31), and this is also likely to be the case for the other related metal resistance systems. This translocation function across the outer membrane is specific for the cation substrate, which explains the lower homology of the C protein, even for operons sharing a common substrate. The membrane orientation of the C protein is believed to be important in the ultimate translocation of the cation from CzcC to an Omp, which would transport the cations away from the cell surface (31). Recently, cation/proton antiporter systems for metal cation efflux have also been described in gram-positives species. Interestingly, these units contain only two components, which are homologues of the A and B proteins of the gramnegatives species and seem to lack the OMF, the C protein. Examples include

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the newly described NrsAB proteins that are involved in nickel, cobalt, and zinc homeostasis of Synechocystis sp. (46). 3.1.2

Efflux Based on Cation Diffusion Facilitation

This resistance mechanism is based on metal cation efflux by an inner membrane protein termed the cation diffusion facilitator, or CDF (39,41,43). CDFs typically act in concert with other cation efflux mechanisms (Table 1), but alone they can confer a low-level resistance phenotype by way of efflux of metal cations across the cytoplasmic membrane resulting in reduced accumulation in the cytoplasm (39,43). The best characterized bacterial CDF is CzcD, which spans the cytoplasmic membrane of R. eutropha CH34 (39) and is implicated in resistance to zinc, cadmium, and cobalt as well as in the regulation of this resistance mechanism—although this has yet to be fully understood (39). CzcD is thought to be functional at very low metal concentrations, where it may act in the uptake of metal cations into the cytoplasm, while at high concentrations of metals, the protein would sponge up metal cations through high-affinity binding by its numerous His residues and exporting the metal substrates across the cytoplasmic membrane (39). It has yet to be determined whether CzcD is itself a regulatory protein or simply an accessory protein whose export activity affects one or more regulators (33,39,47). Eukaryotic homologues of CzcD include COT1 (48) and ZRC1 (49) of Saccharomyces cerevisiae, which efflux cobalt and zinc, respectively, from the cytoplasm to the mitochondria. 3.1.3

Heavy Metal P-Type Efflux ATPases

ATPases are protein pumps that utilize the energy derived from ATP hydrolysis to actively transport substrates from one cellular compartment to another; their enzyme activity is limited to ATP hydrolysis, which means that the substrate is left unchanged. All known ATPases involved in transport of heavy metals for resistance purposes are membrane-spanning, typically integral proteins of the cytoplasmic membrane, and function in the outward translocation of the metals, i.e., expulsion from the cell interior to the exterior. The requirement for ATP hydrolysis taxes the organism for energy, which could explain why many of the metal resistance ATPases function in the elimination of highly toxic metal cations. Some of the best characterized ATPases involved in heavy metal resistance are those belonging to the P-type CadA family (50), which occur ubiquitously in both gram-positive and gram-negative genera. CadA of Staphylococcus aureus, the model system, functions in cadmium, lead, and bismuth efflux (reviewed in 25–27); it achieves this by turning over an aspartyl phosphate intermediate during the catalytic cycle (51,52; reviewed in 53 and 54). Other members of this subfamily include PbrA, functioning specifically for lead efflux in R. eutropha CH34 (55); SilP for silver efflux in S. typhimurium (38); ZntA for zinc, cadmium, and

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lead efflux in Escherichia coli (56,57) and the closely related ZiaA for zinc efflux in Synechocystis (58); and CopF for copper efflux in R. eutropha CH34 (22). The other types of ATPases involved in metal efflux include the chromosomally encoded P-type ATPase CopB for copper export in E. hirae (which acts in concert with the CopA uptake ATPase; 59, reviewed in 60 and 61), ChrD for chromate efflux in R. eutropha CH34 (62), and ArsA of E. coli (reviewed in 63). In R. eutropha CH34, the amino acid sequence of the newly described ChrD protein, which has been implicated in the resistance mechanism to Cr6⫹, shares homology with ATPases of ABC transporters (62). However, the precise role of this ATPase function in relation to the main efflux-based resistance mechanism of the ChrAB proteins (see below) has yet to be demonstrated. The gram-positive ATPases are generally able to pump out the metal cations directly from the cytoplasm, whereas the gram-negative ATPases require additional protein functions to ensure that the cations are translocated across both the inner and outer membranes unidirectionally toward the outside, without leakage into the periplasmic space. Thus, examples of ATPase accessory proteins include ArsB, which complements ArsA for arsenite efflux in E. coli (8,64), and PbrB, a putative outer membrane lipoprotein that likely channels the translocation of Pb2⫹ ions from the PbrA ATPase across the outer membrane of R. eutropha CH34 to the cell exterior (55). 3.1.4

Other Types of Efflux Mechanisms

Copper resistance in both gram-positive and gram-negative bacteria typically involves a combination of uptake, efflux, and, in some cases, storage. Copper is an essential micronutrient for bacteria; its utilization and movement in the cell are intricately controlled because free copper ions can be very toxic. The best characterized resistance determinants include the pco of E. coli (reviewed in 65), the cop operons of Pseudomonas (53,60,66) and R. eutropha CH34 (22,23), while copper tolerance has been characterized for the chromosomally encoded cop operon of E. hirae (59; reviewed in 60,61). The common features of the E. coli, R. eutropha, and Pseudomonas copper resistance systems include the ABCD structural genes that encode an inner membrane uptake protein (PcoD/CopD), an outer membrane uptake protein (PcoB/CopB), and two periplasmic copperbinding proteins (PcoA/CopA and PcoC/CopC). Both the CopA and CopB bind copper with high affinities: for example, the CopB protein of R. eutropha CH34 can potentially bind 23 Cu2⫹ cations per monomer (22). As described in a previous section, R. eutropha CH34 also mediates efflux by the CopF ATPase; however, the physiological interplay between this and the CopABCD system has yet to be defined. Broad host range expression of nickel resistance in proteobacteria is mediated by the NreAB proteins (36,67,68). The mechanism is based on efflux, which is thought to be principally carried out by the NreB protein (68). In R. eutropha

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31A, the Nre system occurs together with the Ncc cation/proton antiporter system, resulting in the expression of high-level nickel resistance (36,68). In Synechocystis sp., the NrsD protein, which shows high homology with the NreB protein of R. eutropha 31A (46), was shown to be a membrane-spanning divalent cation transport protein of the major facilitator superfamily (reviewed in 69) that binds nickel with high affinity. 3.1.5

Enzymatic Reduction

Enzymatic reduction is a highly specific resistance mechanism adopted by bacteria for the transformation and removal of highly toxic metal species. Enzymatic reduction of metal cations is highly energy-dependent, and in some examples of resistance, chemical reduction can result in a metal species that is more toxic. The best characterized is mercury resistance, which is also the single most widespread of all antimicrobial resistances. The Tn501-localized mer operon of E. coli is perhaps the best studied of all metal resistance systems (see review in 70). Mercuric ions are extremely toxic to bacteria because they can interact nonselectively with the thiol groups of proteins. Therefore, a complex mechanism of resistance has evolved, which has been relatively well conserved among eubacteria (25,26). The mechanism involves uptake of Hg2⫹ into the cell interior to avoid its deleterious interaction with cellular components, and this is carried out first by the MerP protein in the periplasm and then the MerT protein that spans the inner membrane (70). The mercuric reductase protein, MerA, which is localized in the cytoplasm, enzymatically reduces the toxic Hg2⫹ ions to Hg0: this nontoxic end product simply volatizes out of the cell. Mercuric reductase is a highly specific metalloenzyme reacting only with Hg2⫹; therefore, in bacteria harboring a broad-spectrum resistance to inorganic Hg2⫹ as well as organomercurial compounds the reductase functions in concert with an organomercurial lyase enzyme, MerB, which cleaves the Hg2⫹ moiety prior to chemical reduction by the MerA protein (32,71,72). Arsenate reductase, or ArsC, is the main component of the plasmid-encoded arsenate resistance mechanism of S. aureus (27,73,74), E. coli (75), and chromosomally encoded mechanisms of some Pseudomonas species (76) and E. coli (77). This cytoplasmic metalloreductase reduces intracellular arsenate (As5⫹) to the more toxic arsenite (As3⫹). The latter is extremely reactive with the thiol groups of proteins and is therefore rapidly expunged by the ArsB protein, which in S. aureus transports the arsenite across the cell membrane in an ATPaseindependent manner, while in E. coli and other gram-negatives bacteria the ArsA and ArsB proteins combine to form an efflux ATPase (64,78; reviewed in 27). Metabolically, it seems paradoxical for the organism to intentionally convert the already toxic As5⫹ to the more toxic As3⫹. However, the reasoning may be related to maintaining phosphate homeostasis; since arsenate is structurally similar to the phosphate oxyanion, the organism alters the contaminating metal to arsenite,

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which has no analogue that is essential to metabolism, and can therefore be rapidly removed from the cell. A Desulfomicrobium strain has been described that utilizes the reduction of arsenate to arsenite to meet respiration requirements rather than as a resistance mechanism (79). Arsenate reduction is the more widespread resistance mechanism for arsenicals; however, some bacteria confer resistance by oxidizing arsenite and other arsenocompunds to the less toxic arsenate, with the most effective arsenite oxidizer being Alcaligenes faecalis (80). Also, there are two recent, independent reports of new types of chemolithoautotrophic gram-negative species that are capable of oxidizing arsenite to arsenate (81,82), which in the strain NT-26 is effected by the periplasmic-located arsenite oxidase (82). Bacteria have also been described that can oxidize arsenite to elemental arsenic (reviewed in 27). Chromate resistance has been shown to include a reduction of the toxic Cr6⫹ cation to the less toxic and less soluble Cr3⫹. As yet, no specific chromate reductase has been linked to chromate resistance, and it is believed that a chromosomal reductase—most likely one involved in sulfate metabolism—carries out this function in a variety of bacteria (62,83–87). 3.1.6

Other Mechanisms of Heavy Metal Resistance

Described below are mechanisms that in themselves are unique due to the specific toxicities or cellular utilization of the metal, or simply because they have only recently been described and await detailed characterization. For the toxic metals, a recurring theme of the resistance mechanism is cation uptake, immediately followed by transformation and then export from the cell. Specific examples are given below. In the case of chromate, which is considered to be highly toxic to biological systems because it can interfere with sulfate transport systems, both plasmid and chromosomal determinants have been reported. The ChrA protein is an as-yet uncharacterized transmembrane efflux protein that functions in the expulsion of the toxic chromate in P. aeruginosa (84), Bacillus spp. and Enterobacter cloacae (86) and R. eutropha. Chromate is capable of entering the cell interior via the sulfate uptake systems. Under sulfate-starved conditions, chromate is readily taken up into the cell and rapidly reduced by the chromosomally encoded sulfate reduction systems (87). The rapid turnover of Cr(VI) to the less soluble Cr(III) (which is not an energetically favorable enzymatic process for the bacteria) does not allow for the induction of the chr genes, which are activated by Chr(VI) (87). However, under sulfate-saturated conditions, chromate uptake into the cells occurs more slowly. Concomitantly, the sulfate-reducing systems become inactivated and this results in the intracellular accumulation of Cr(VI)—the trigger for induction of chr expression. In R. eutropha, ChrA activity has been shown to be induced under sulfate-saturated conditions (87). Thus, ChrA activity is inducible on accumulation of intracellular Cr(VI) which occurs under sulfate-saturated con-

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ditions, and this triggers a rapid expulsion of the toxic oxy anion. The function of the ChrB protein, which is also essential for chromate resistance, has yet to be defined (30,62), whereas the ChrC protein shows homology with a manganesecontaining superoxide dismutase and ChrD is likely to be an ATPase homologous to those of ABC transporters. 3.1.7

Complexation, Sequestration, and Precipitation of Heavy Metals

Sequestration of heavy metals in different compartments of the bacterial cell can serve two purposes in dealing with high concentrations of heavy metals: (a) to immobilize the metal to prevent it from moving freely within the cell, which can be reversible or not, and (b) to concentrate or precipitate the metal out of solution irreversibly in order to completely abolish its toxic activity. The cellular sequestration of metals that occurs in concert with resistance mechanisms largely depends on the metal species itself, on how the metal is exported out of the cell, and on the metal concentration. Sequestration of the metal typically involves high-affinity proteins as well as nonspecific interactions with polymeric substances or other types of biological adsorbents produced by the cell. 3.1.7.1 Extracellular. In CH34, postefflux sequestration of the metal cations at the cell surface is an important mechanism that prevents reentry of the metal into the cell, particularly when the extracellular concentrations are high (32,88,89). In cultures of CH34, zinc and cadmium removal results in immobilization of the metal cations as complexes with carbonates, bicarbonates, and hydroxides. These processes were found to be induced by carbon dioxide emitted by the metabolic activity as well as alkalinization of the extracellular environment resulting from the cation/proton antiporter activity, while in other R. eutropha strains secretion of extracellular polysaccharides has also been implicated. These metal complexes are precipitated as crystals at the cell surface (21,32,88,89). This bioprecipitation process precedes the release of extracellular polysaccharides and nucleation proteins, which also act to sequester the metal at the cell surface, as demonstrated for the R. eutropha-like strain ER121 (42). As described later in this chapter, the physical immobilization of the metals outside the bacterial cells has been exploited in bioreactor designs for the removal and recovery of metals from metal-contaminated wastes (32,90). It was recently demonstrated that cnr- and ncc-mediated resistance mechanisms also result in metal sequestration, which in accumulation experiments with nickel could result in 48% removal of the metal from culture supernatants (S. Taghavi, personal communication). Furthermore, when CH34 is pregrown in the presence of zinc or cadmium and transferred to a medium with high levels of nickel, metal precipitation at the cell surface is observed (32). Thus, a common postefflux sequestration/immobilization mechanism seems to be operable in CH34; the pre-

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cise nature of this process has yet to be pinpointed, although precipitation kinetics implicate high-pH-induced complexation in the form of metal carbonates as being a major underlying cause (32,88,89,90). It follows that because cation/proton antiporter efflux mechanisms by their nature lead to high localized concentrations of metal at the cell surface, then all CBA-type mechanisms can potentially induce some type of postefflux immobilization of their substrates. Some gram-negative organisms have evolved resistance mechanisms that allow for the sequestration of high concentrations of heavy metals in the periplasmic space. For essential cations, such as copper, a newly assigned gene is copH whose gene product shares high homology with CzcE of R. eutropha CH34, which may function in cadmium, zinc, and cobalt sequestration in the periplasm (van der Lelie et al., unpublished data). The function of CopH is still under investigation; however, preliminary data indicate that CopH is essential for copper resistance since copH mutants displayed a reduced copper resistance phenotype (van der Lelie, unpublished data). In P. syringae var. tomato, excess copper is stored in the periplasmic space in its blue form (Cu2⫹) by an unknown mechanism that may involve high-affinity copper-binding proteins (60). In E. coli containing plasmid pRJ1004, no periplasmic copper storage has been observed (91), although an additional periplasmic protein, PcoE, has been described which is a high-affinity copper-binding protein that may be implicated in this function (92). PcoE is homologous to the silverbinding periplasmic SilE protein, whose crystal structure was recently solved (38). The SilE protein structure is novel: a periplasmic protein, SilE, binds four Ag⫹ cations per monomer. Apo-SilE has very little secondary structure that likely maintains it in an inactive state, while metallized SilE adopts a highly ordered structure rich in α-helical content, resulting in protein activation (Silver, personal communication). This type of structural control on protein activity prevents fortuitous activation by cations other than Ag⫹, which renders the system highly specific for silver cations. 3.1.7.2 Intracellular Intracellular sequestration of heavy metals is a mechanism shared between prokaryotic and eukaryotic systems. This is due to a family of small proteins called metallothioneins that are typically short polypeptides of approximately 60 residues. These proteins are rich in cysteine residues and bind heavy metals with high affinity, thereby immobilizing them within the cytoplasm (reviewed in 27). Metallothioneins are widely dispersed in eukaryotes, the best characterized being that of the Norwegian rat, Rattus norvegicus (93), while the best characterized for bacteria is SmtA which effects zinc resistance in the cyanobacterium Synechococcus (94). Most eukaryotic metallothioneins bind nonspecifically to certain families of metal species, whereas SmtA binds with high affinity to its substrates Zn2⫹ and Cd2⫹ (with a preference for the former)

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in metal-binding clusters, which are organized at the N and C termini of the protein (94). Lead resistance in S. aureus and Citrobacter freundii has also been shown to involve the detoxification of Pb2⫹ cations to a less reactive species and the subsequent intracellular immobilization of the metal as lead-phosphate complexes (95). This prevents the detoxified lead from participating in any cellular activity that could be potentially harmful to the organism and also permits the organism from wasting energy in pumping out the cation. The recently described lead resistance operon of R. eutropha CD34 may also mediate sequestration of the metal, and PbrD, a presumed Pb-binding protein, is believed to be involved in this function since deletion mutants abrogated lead sequestration in CH34 derivative strains (55). 3.2

Physiological Role of the Metal Determines the Resistance Mechanism

Resistance to heavy metals that are required as trace elements for bacterial metabolism poses an interesting paradox for bacteria. On the one hand, resistance against toxic levels of the metal confers a selective advantage of the bacteria as compared with sensitive commensals within a given biotope; however, the bacteria must ensure that resistance does not jeopardize the normal homeostatic utilization of the metal required for normal growth. The evolution of bacterial resistance and tolerance to metals has been widely debated during the past decade for several reasons: (a) there seems to be common features in the driving forces by which bacteria recruit heavy metal and drug resistance genes; (b) the evolutionary choice of the mechanism of resistance that the bacteria adopts seems to be closely related to whether the metal is essential or toxic (reviewed in 28). The essential metals include zinc, nickel, cobalt, and copper, while the highly toxic metals include cadmium, mercury, silver, arsenic, chromate, thallium, lead, strontium, and bismuth. Most resistance mechanisms are inducible by metal, which enables the organism to expend energy only when absolutely necessary. Thus, in some regulation mutant strains, such as the R. eutropha CH34 derivative AE963, which constitutively expresses the cnr operon and as a result becomes nickel-deficient (96), the organism can become dependent on the metal presence for normal growth. 3.2.1

Regulation of Heavy Metal Resistance

As a general principle, bacterial resistance mechanisms must necessarily take into account the metabolic requirements of the organism and the physiological role of the metal. A finely tuned regulation of the resistance mechanisms is therefore a prerequisite so as not to jeopardize homeostasis.

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As suggested by Table 1, most metal resistance mechanisms can be generally divided into those based on efflux, enzymatic conversion, or sequestration. The majority of the coding regions (structural genes) seem to exist as modules (Fig. 1). However, the genes of regulation seem to have been acquired by a mix and matching of different genetic systems, which may have taken place at different evolutionary stages (Table 2) (reviewed in 1,97–99). 3.2.1.1 Regulation of Cation/Proton Antiporters As shown in Table 2, the main types of regulatory modules controlling cation/proton antiporter-based resistance systems are the two-component responder/sensor units. Examples include PcoRS and CopRS, controlling copper resistance of E. coli and Pseudomonas, respectively (reviewed in 60); CzrRS, controlling cadmium and zinc resistance in P. aeruginosa (37); and SilRS, controlling silver resistance in S. typhimurium (38). The sensor or ‘‘S’’ protein is a histidine kinase that autophosphorylates itself at a His residue on interaction with its metal substrate and is typically membrane spanning. Autophosphorylation of the S protein causes an allosteric conformational change in the C terminus of the protein that is transduced across the inner membrane to the cytoplasmic responder protein. The latter is activated by phosphorylation of an Asp residue by the kinase activity of the S protein. In the inactive state, the responder will typically act as DNA-binding repressor, preventing transcription of the structural genes. The responder protein is activated by phosphorylation, which relieves DNA repression by abrogating the DNA binding activity. In general, all two-component sensor-responder systems constitute a highly efficient on-and-off switch for expression of resistance, allowing a fine-tuning of the level of induction in proportion to the metal concentration. This is certainly true for the czc operon of R. eutropha CH34, where regulation by the two-component CzcRS allows the bacteria to respond optimally in a micromolar to millimolar concentration range of Zn2⫹ (87). The lower micromolar threshold allows CH34 to take up zinc freely for growth requirements, and the effective working range of the regulatory unit allows the organism to modulate efflux of excess metal in proportion to the extracellular levels. CH34 also

FIG. 1 Genetic relationships between bacterial metal resistance mechanisms. Boxes and lines indicate the relationships based on amino acid sequence homology. Vertical lines indicate homology between blocks of genes. The resistance operons are as follows: cnr (cobalt and nickel), czc (cadmium, zinc, cobalt), ncc (nickel, cobalt, and cadmium), czr (cadmium and zinc), sil (silver), cop, pco (copper), chr (chromate), cad (cadmium), zia, znt (zinc), ars (arsenate), pbr (lead), mer (mercury). mgt encodes a presumed Mg2⫹ transport protein. ‘‘C’’ and ‘‘P’’ indicate chromosomal and plasmid-encoded, respectively.

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20

TABLE 2 Best-Studied Regulatory Modules Involved in Regulation of Heavy Metal Resistance Mechanisms in Bacteria Structural Resistance

Cation/proton antiporter

MerR transcriptional regulators

Two-component responder-sensors

ATPase

Enzymatic reduction

ZntR CoaR PbrR

Sequestration

PbrR

CzcRS CzrRS SilRS CopRS CopRS PcoRS

ECF-Based ArsR transcriptional regulators

CadC ZiaR ArsR SmtB

Others

CzcNI ArsD MerD

138,104 105 55 33,47 37 38 22,60 60 60 35,100 36 112 58 110 114 22,47 119 108

van der Lelie and Tibazarwa

CnrYXH NccYXH

Ref.

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makes use of the additional functions of CzcD (and perhaps CzcN and Czcl), which are thought to control the upper limits of transcription of the resistance genes to prevent excessive metal efflux (47,87). For czc, which effluxes zinc, cadmium, and cobalt, regulation is chiefly controlled by the zinc concentration, which is the most potent inducer of CzcRS activity, and this allows the organism to modulate expression of resistance without affecting the homeostasis. Interestingly, the cation proton antiporter systems encoded by the cnr and ncc operons of R. eutropha strains CH34 and 31A, respectively, are controlled by a newly described three-component regulatory unit, based on alternative σ factors. The Ncc and Cnr systems are closely related; both are essentially nickel efflux systems and are therefore mainly regulated by nickel (35,96,100). The regulation of cnr was recently described and seems to involve an intricate interplay between CnrH, an alternative σ factor of the ECF family; CnrY, a putative anti-σ factor that antagonizes CnrH activity; and CnrX, a periplasmic protein that may function as a metal sensor (100,101). The CnrH protein is a DNAbinding activator protein that specifically directs transcription of the cnr genes in the presence of RNA polymerase (100,102). The activity is controlled at the protein level by the CnrY and CnrX proteins, which constitute the metal-sensing unit. CnrH is capable of activating cnr transcription up to 1000-fold between 100 to 300 µM of nickel, and can modulate transcription activation as a function of both nickel and cobalt concentrations in the micromolar to millimolar range (100,101). This ECF-based regulation has yet to be described for other metal resistance operons, although CnrH has homologues, including RpoE (or SigE), which also controls operons involved in extracellular sensing and stress responses (reviewed in 102,103). 3.2.1.2 Regulation of ATPases As shown in Table 2, three main families of regulators have been described and include the MerR family and ArsR subfamilies of transcriptional regulators as well as the two-component responder/sensor units. Both the MerR and ArsR classes are metal-binding, DNA-binding proteins, which repress gene transcription in the absence of inducer metal and whose function is abrogated by the specific binding of their cognate metal cation ligand. The MerR family of proteins includes MerR, which controls the merencoded mercury resistance found ubiquitously in eubateria; ZntR, which controls the zinc resistance, znt operon of E. coli (104); CoaR, controlling cobalt resistance in Synechocystis (105); ZiaR, controlling the ZiaA zinc ATPase of Synechocystis sp. (58); and PbrR, controlling the pbr operon encoding lead resistance in R. eutropha CH34 (55). Only MerR, PbrR, ZntR, and CoaR have been characterized at the protein level. MerR is a small, dimeric protein with an N-terminal DNA binding domain and a mercury binding site at its C terminus. In the absence of inducer Hg2⫹, MerR binds to an operator site (a region of dyad symmetry), which lies between

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the ⫺35 and ⫺10 regions of the merT promoter (PT) and actively represses RNA polymerase holoenzyme from transcribing merR or the structural genes, thereby actively repressing all mer genes, including its own. Activation of the mer operon is effected by the binding of a single Hg2⫹ to the MerR dimer (106); structural changes that result from the binding of the Hg-MerR complex to its cognate PT promoter cause a localized distortion of the DNA in the promoter region thereby allowing the ⫺35 and ⫺10 regions to align adequately as to be recognized by core RNA polymerase. The distortion is detrimental to the structural integrity of the merR promoter, PR, so that it becomes inactive (107). Downregulation of mer expression at low mercuric ion concentrations is effected by the MerD protein, which can bind the operator region in PT, resulting in reduced transcription from this promoter (108). This mechanism of transcriptional control is one of the best examples of tight regulation known in eubacteria. ZntR has only recently been identified and characterized (104). Like MerR, ZntR activates transcription from PzntA, the promoter of zntA; both Cd2⫹ and Hg2⫹ have a low induction effect (in the micromolar range) on transcription from PzntA. Nevertheless, induction of transcription from this promoter was observed to reach a maximum of about 100-fold induction in the presence of 1 mM Zn2⫹, compared with 100,000-fold induction resulting from MerR activation from the Pmer promoter (70). One reason for this lies in the observation that ZntR does not actively repress transcription from its cognate promoter as does MerR (104), which in the absence of inducer Hg2⫹ represses transcription up to 10-fold (65). This difference in promoter strengths between the mer and znt operons reflects the importance of maintaining the homeostatic balance of Zn2⫹ compared with the need to expulse toxic levels of mercury. The ArsR subfamily of repressors includes ArsR, which controls expression of the ars operon encoding arsenate resistance in a variety of eubacteria (109– 111); CadC, which regulates the cad operon encoding resistance to cadmium, zinc, and lead in a large number of gram-positive and gram-negative bacteria (57,112,113); and SmtB, whose crystal structure was recently resolved and which regulates the expression of the SmtA metallothionein of Synechococcus sp. (94,114). This group of proteins are homodimeric DNA- and metal-binding proteins, that have a unique pair of cysteines near the C-terminal region, which is implicated in both DNA binding and cation sensing (25,26,112,114–117). In the absence of inducer metal, the repressors bind to an operator region downstream of their own promoter, thereby inhibiting transcription of the structural genes as well as their own transcription. In the case of CadC, the repressor has a high affinity for DNA; however, its interaction with metal cations, including Mn2⫹, Zn2⫹, Cd2⫹, Bi2⫹, and Pb2⫹, cause the repressor to dissociate from the DNA molecule, thereby allowing transcription of the resistance genes (112,113). This allows CadA ATPase to respond to a variety of potentially toxic metal cations. The

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crystal structure of SmtB was recently solved and found to be a homodimeric protein (61,114). Two-component systems are also involved in the regulation of copper resistance ATPases and include CopRS, which might regulate the CopF ATPase of R. eutropha CH34 (22). The CopYZ unit that regulates the cop operon of E. hirae is also two-component, although it is not a classical responder/sensor system. CopY is a DNA-binding repressor protein and CopZ an antirepressor protein (118). This regulatory mechanism, which is unique for a metal resistance operon, is effected by the CopY aporepressor, which cannot bind to its cognate operator/ promoter DNA in the absence of Cu⫹ (the preferred substrate in vitro). However, at physiologically tolerable levels of the inducer metal cation, the complex (CopY-Cu⫹)2 is converted to a DNA-binding repressor. At higher metal concentrations, the antirepressor CopZ binds Cu⫹ and the CopZ-Cu⫹ complex interacts with (CopY-Cu⫹)2 to form an inhibitory complex that disrupts the DNA binding activity of the latter. This regulatory mechanism allows a fine-tuned response to extracellular copper, since it allows the expression of uptake proteins at low concentrations of the metal and the expression of the efflux proteins when physiological tolerable thresholds are surpassed. Thus, metal homeostasis and resistance mechanisms are simultaneously controlled (118). 3.2.1.3 Regulation of Enzymatic Reduction As mentioned above, enzymatic reduction is an effective resistance mechanism for the detoxification of highly toxic metal cations. Therefore, most regulators of these mechanisms tend to be highly cation-specific transcriptional regulators. Examples include ArsR, controlling expression of the ArsC arsenate reductase (109–111), and MerR, controlling expression of the MerA mercuric reductase (70). Regulation of expression of the enzymatic reduction processes seems to involve an additional regulator that limits the maximum expression levels of resistance above the control levels of the main transcriptional regulators. Examples include the DNA-binding repressors ArsD (119,120) and MerD (108), which set the upper limits of expression of the ars and mer operons, respectively. 3.2.1.4 Regulation of Metal Sequestration Processes In most cases, sequestration is a complex process involving proteins, polysaccharides, and so forth; therefore, the underlying mechanisms of control are not always well defined. In the case of copper sequestration, the Pco and Cop systems of gramnegative species are likely to be under the control of the two-component responder/sensor regulatory units that respond specificically to Cu cations and that control expression of the resistance gene (see above). Sequestration of the SmtA metallothionein is controlled by the SmtB protein (see above), which responds to both Zn2⫹ and Cd2⫹.

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APPLICATIONS BASED ON BACTERIAL INTERACTIONS WITH HEAVY METALS

4.1

Metal Leaching

Metal leaching is based on metabolic activity of the bacteria in the presence of heavy metals. Metal leaching activities for the purposes of mining from metal ores has been recently reviewed (5). The key organisms implicated include members of the Thiobacillus and Leptospirillum genera. Thiobacillus ferrooxidans is the most frequently used and, as described earlier, this is due to its interesting metabolism that allows it to oxidize ferrous iron and other heavy metals as part of a sulfur cycle required by the organism for its energy production. Under strict acidic conditions, this results in the production of soluble metal sulfates (reviewed in 5, 10, 11), a process also known as ‘‘leaching’’ (12). This process occurs naturally in iron-rich sulfide deposits. However, the presence of heavy metal impurities, including nickel, copper, lead, gallium, zinc, cobalt, and copper, has been shown to cause the release of high concentrations of soluble heavy metals. This process is used at an industrial scale for leaching of metals from low-grade ores. This is typically carried out in heaps (‘‘heap’’ or ‘‘dump’’ leaching), where the ore is stacked as a heap on a slope or inclination, and acidified water is poured over this heap to lower the pH and to stimulate the activity of the acidophilic bacteria. The aqueous metal fraction that leaches out of the ore is collected at the bottom of the slope. Similar principles have been applied to construct continuous-flow bioreactors where inoculants of T. ferroxidans were used to leach out contaminating metals from sewage sludge resulting in the solubilization of more than 90% of the otherwise insoluble metal fraction, thereby decontaminating the sludge for safe disposal (121,122). Leaching of heavy metals from solid contaminated wastes has also been described based on citric acid excretion by P. putida, which could effectively solubilize zinc from industrial filter dust (123). 4.2

Decontamination of Soils and Water

Biological treatment methods for decontamination of metal-polluted soils have been given increasing attention since traditional remediation techniques are often ineffective at desorbing the metal contaminants from soil particles. 4.2.1

Applications Based on Ralstonia eutropha CH34 and Related Strains

The interaction of CH34 with metals is a feature that has been exploited in the development of various technologies for environmental applications; these have been widely documented and were recently reviewed in 9,22, and 23). Some applications are based on the function of the metal resistance structural proteins

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and metal precipitation; others exploit the capability of heavy metals to control gene activity in CH34. In addition, CH34 is capable of degrading certain aromatic xenobiotic compounds, including biphenyl (124), which has extended the range of applications of this organism for the treatment of a wide range of industrial wastes containing mixed pollution (125). 4.2.1.1 Heavy Metal Removal and Recovery from Contaminated Effluents One of the more successful technologies developed is the BICMER, or Bacteria Immobilized Composite MEmbrane Reactor (42,88,89,125). The design of the BICMER is founded on the underlying mechanism of postefflux metal precipitation by CH34. This novel type of bioreactor consists of a specially designed Zirfon membrane on whose surface the active bacteria are immobilised, thus forming a biofilm. The waste stream contaminated with heavy metals and/or organic xenobiotics contacts the bacteria within a tubular structure or a flat sheet support while a separated feeding stream keeps the bacteria viable. CH34 activity results in metal precipitation as insoluble crystals on the microbial biofilm that are subsequently recovered on glass bead columns. This bioreactor concept has been successfully applied for the removal and recovery of various heavy metals, including Zn, Cd, Co, Ni, Cu Pb, Y, and Ge, from wastewaters or process streams. The BICMER has many advantages over other related technologies, including the following: (a) the concentration effect of the pollutant on the membrane reduces the sludge volume; (b) the insoluble metal carbonate crystals allow direct recovery of the metal; and (c) the ability of CH34 to degrade certain organic compounds means that the biofilm can tolerate wastewater loads containing recalcitrant organic compounds, including polyaromatic hydrocarbons (PAHs) (125). A similar concept based on the biosorption and bioprecipitation properties of the Ralstonia biofilms has been applied in the development of a moving-bed sand filter where the biofilm is allowed to form on sand grains (90). The resulting metal-loaded biomass, i.e., bacteria containing immobilized metals even up to 10% of dry matter, is recovered and collected. These sand filters are being developed as a ‘‘polishing’’ technology for treatment of heavily contaminated effluents (90). 4.2.1.2 Soil Bioremediation Two key features of CH34 physiology have been exploited in the development of the BMSR or Bio Metal Sludge Reactor for contaminated soils, namely, that the organism can sequester heavy metals and adsorb them onto its outer surface and that it can also stimulate soil settling, including the fraction of fine organic matter, while the heavy metal–loaded bacteria remain in suspension. The BMSR concept involves a direct inoculation of the soil with the metal-resistant bacteria into a stirred reactor. Interestingly, CH34 activity alters the colloidal properties of the treated soil by improving its settling capacity, thus allowing the treated soil to be separated and recovered from the heavy metal–loaded biomass consisting of bacteria charged with heavy metals

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which remain in suspension. This biomass can subsequently be removed from the process water. The heavy metal content of the resulting biomass can attain 0.5% on a weight-for-weight basis (90) while the overall performance of the BMSR can result in up to 85% of the bioavailable metals from the soils to be extracted within 48 h. 4.2.1.3 Metal Immobilization Using Biomolecules This is based on the immobilization of the metals in the soil using biopolymers or biosurfactants. Biopolymers bind metals with high affinity and can be physically separated from waste effluents or cleansed soils, thus allowing metal recovery. Examples include exopolysaccharides, cyclodextrins, and biosurfactants. These biopolymers complex divalent metal cations and render them more soluble in aqueous solution, which can be further exploited in soil washing techniques for maximal removal of the metal contaminants in the soil (reviewed in 126). 4.2.2

Environmental Monitoring and Metal Sensing: Heavy Metal Biosensors

The application of bacterial resistance mechanisms to develop heavy metal– sensing biological tools has been widely explored and is now reaching full potential. With the recent advances made in our understanding of the bacterial heavy metal resistance mechanisms, it has been possible to engineer bacteria that produce a measurable and quantitative signal when brought in contact with specific metal species (127–131). 4.2.2.1 Gene-Based Biosensors To date, the most successful metal biosensors are the gene-based biosensors constructed by the fusion of metal resistance regulators with the bioluminescent luciferase genes of Vibrio spp. (127,129,132,133) and of Photobacterium phosphoreum (reviewed in 134). The most important criterion for such transcriptional fusions is a metal-responsive promoter, which ideally should be significantly activated in the presence of metal. Transcriptional fusions incorporating both a metal-inducible promoter and a gene encoding a regulator can allow for a fine-tuned transcriptional response by the biosensor construct on activation by metals and these (reviewed in 129). Promoterless vector plasmids have been constructed based on the luciferase reporter genes, luxCDABE, and these have facilitated the cloning of metal-responsive genes so that in biosensor constructs light emission occurs as a function of metalinduced transcription (128,129,131,133). Using different regulatory regions of CH34 resistance determinants, it has been possible to develop an amalgam of BIOMET (biological metal) sensors for different metals and families of metals (23,131). These are summarized in Table 3. The gene-based biosensors have a unique niche in environmental monitoring applications since they are one of the few techniques that can give information on the bioavailability of a metal species in a particular sample (129,130). Quanti-

Bacteria

TABLE 3 Gene-Based BIOMET Biosensor Strains Developed in R. eutropha CH34

Strain AE1239 AE1433 AE1433 AE2440 AE2450 AE2515

Cation specificity

Metalresponsive genes

Detection limit (mg/kg dw)

Industrial soil/ sediment (mg/kg dw)

Background soil (mg/kg dw)

Ref.

Cu2⫹ Zn2⫹ Cd2⫹ Cr6⫹ Pb2⫹ Ni2⫹

cop czc czc chr pbr cnr

18.7 5.9 6.37 1.21 0.83 2.2

673–1365 2406–7159 18–109 683–1319 2192–5788 538–1731

14–29 50–148 0.5–2.9 32–61 35–93 7–22

128 133 133 131 131 135

a

All constructions are based on the bioluminescent lux reporter genes of Vibrio fischeri. The biosensor strains have been standardized for their response to specific metal cations in soil and other environmental samples. The detection limit indicates the lowest metal concentration for a quantifiable biosensor response. Also included in the table are data on the typical contamination levels in industrial soils/sediments as well as the norms in noncontaminated soils.

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tation of the bioavailability of a metal species is perhaps the most important parameter for assessing the relative toxicity of a sample to biological systems (130). The BIOMET tests were found to be a determining criterion in the evaluation of in situ inactivation of metal-contaminated soils following treatment with chemical additives (17); the usefulness of these tests was maximally realized when used in conjunction with chemical and biological tests based on phytoxicity, zootoxicity, and bacterial ecology (135). Other applications of the BIOMET sensors include the routine monitoring of bioavailable metal levels in soils subjected to remediation strategies (23,90) and measurement of metal content in industrial wastes/byproducts, such as incinerator fly ashes (131,133). 4.2.2.2 Protein-Based Biosensors The next generation of biosensors will likely be the protein-based biosensors that are currently being developed alongside the gene-based variety (131). These sensors are based on the immobilization of metal-responsive proteins (typically metal-binding metalloregulatory proteins) or other metal-binding proteins onto electrodes; the sensing quality is measured as a change in capacitance when the protein interacts with the metal (131). Biosensors based on MerR (70,136) have been successfully applied for sensing of low Hg2⫹ concentrations, whereas a recombinant form of SmtA could sense Cu2⫹ levels between 10⫺5and 10⫺1 M (131).

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2 Mycorrhizal Fungi Jan V. Colpaert Limburgs Universitair Centrum, Diepenbeek, Belgium

P. Vandenkoornhuyse Universite´ de Lille, Paris, France

1

INTRODUCTION

Mycorrhizal symbiosis is the most widespread symbiosis between plants and microorganisms. Between 80% and 90% of all seed plant species have fungi in their roots, forming structures known as mycorrhizas (1). Mycorrhizas are a functional part of the plant roots where the fungus provides an interface between the roots and the soil. Mycorrhizas are very effective in assimilating nutrients, including essential metals and their analogues, many of which are present in toxic concentrations on contaminated soils. Mycorrhizas range widely in form and in type of fungus involved, demonstrating that they represent not a single class of symbiosis but rather a type of plant-fungus association that has evolved repeatedly, in response to distinct selection pressures (2). Mycorrhizas have been classified in various ways, such as into ectomycorrhizas, typically formed between some longlived woody plant species and long-lived fungi (generally Basidiomycota and Ascomycota) (Fig. 1), and endomycorrhizas. The latter contain the most ubiquitous mycorrhizal association, the arbuscular mycorrhizas (Fig. 1), involving a 37

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FIG. 1 Longitudinal section of an ectomycorrhiza and an arbuscular mycorrhiza. (Ref. 105.)

very wide range of plants and a small group of fungi in the Glomales (Zygomycota) (3). A much smaller group of plants form the ericoid endomycorrhizas. In the following paragraphs, we describe the three types of mycorrhizas that are commonly found in metal-polluted habitats. Arbuscular mycorrhizal (AM) fungi are strict biotrophs. In exchange for carbon compounds, these symbiotic fungi improve the uptake of phosphorus in plants (1,4,5), a highly immobile element in soil. In some circumstances, AM fungi can also benefit plants by increasing the uptake of nitrogen (4,6) and micronutrients (5,7). In general, AM fungi have a positive effect on plant health and resistance to stress factors (8). The ectomycorrhizal (ECM) fungi are typically associated with woody plant species in boreal and temperate forest ecosystems. Most members of the important tree families (Pinaceae, Fagaceae, Betulaceae, Salicaceae, Tiliaceae) in these climatic regions are predominantly ECM, although species of the (sub)tropical families Myrtaceae and Dipterocarpaceae also appear to be mainly ECM.

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In contrast to the situation found in the AM symbiosis, a large number of fungi have been reported as forming ectomycorrhizas (9). The majority of these species belong to the Basidiomycota, but a significant proportion of Ascomycota are also involved, as are a few zygomycetous fungi. Most ECM fungi are biotrophs in their natural habitats, although some species can be cultured in vitro. They increase the fitness of their host trees in very similar ways as the AM fungi of herbaceous plants. However, ECM fungi seem to be better adapted to organic soil layers in forests, and it is generally accepted that they have better access to organic P and N compounds than plant roots or AM fungi (1). The third distinct form of mycorrhizas can be recognized in members of the Ericales: the ericoid mycorrhizal (ERM) fungi. The very fine hair roots of these plant species are usually intracellularly colonized by a limited (?) number of fungi, mainly ascomycetes belonging to the genera Hymenoscyphus and Oidiodendron. Ericaceous plants are most strongly associated with nutrient-poor soils in which the major growth-limiting nutrient is nitrogen. In the Arctic regions these plants become dominant in soils where a recalcitrant litter accumulates as raw humus. The ERM mycobionts play a key role in the nitrogen nutrition of the plants, and it is thought that the fungi protect the plants against the high metal concentrations (Al, Fe, Mn) present in many habitats of the Ericaceae (1).

2

DO HEAVY METALS IMPOSE A SELECTION PRESSURE ON MYCORRHIZAL COMMUNITIES?

Heavy metals in soils are known to be toxic to most organisms, including plants (this volume, next chapters, 10), when present in excessive concentrations. However, metal tolerances in mycorrhizal fungi have been investigated in much less detail than in their hosts. This is despite the fact that they may be key to plant survival on contaminated soils, not only for potential conferred or enhanced metal resistances but for their role in nutrient and water acquisition as well (11–13). The impact of metal toxicity on vegetations can easily be demonstrated and there is ample evidence for the evolution of adaptive metal tolerance in some higher plants, e.g., in grass species. Other plants, typical metallophytes, are only found on metal-contaminated soil (14). Similar evidence is far less convincing when mycorrhizal communities or even soil-borne microorganisms in general are considered. An important reason for this lack of knowledge is the obvious practical difficulties in assessing the active below-ground microbial communities and populations (15). Recent developments in molecular ecology now provide us with some tools that allow for the detection and accurate identification of belowground fungal partners (16). The accuracy and feasibility of mycorrhizal community studies have been greatly enhanced.

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Although there is evidence that some microorganisms are more sensitive to heavy metal stress than plants growing on the same soil (17,18), heavy metal pollution-mediated changes in plant and fungal communities have seldom been studied in detail on the same site. In many fungal groups, there is little evidence for intraspecific adaptive changes between populations from polluted or nonpolluted habitats (19). Reviewing the effect of metals on soil microbial communities, Duxbury (20) concluded that there is rarely a correlation between the metal tolerance of microorganisms under laboratory conditions and the metal concentrations at the site of origin, unless extremely high metal concentrations are present in the soil. Many of the soil-borne fungi that occur on polluted soils are supposed to have a constitutive high level of metal tolerance, so that selection pressure for more ‘‘adaptive’’ tolerance is seldom exhibited (19,21). Evidence for the evolutionary adaptation to heavy metals among mycorrhizal fungi is scarce. In many cases, heavy metal pollution did not result in an increase in metal tolerance of mycorrhizal fungi, although it is likely that there is selection for those species that possess a high constitutive metal tolerance (21,22). Nevertheless, adverse effects on soil microbial parameters (e.g., specific microbial respiration rate, N2 fixation) and on mycorrhizal communities can be observed in metal-polluted environments (18,23). Such changes might have significant implications for ecosystem functioning and sustainability.

3 3.1

MYCORRHIZAL FUNGI IN HEAVY METAL POLLUTED SOILS Arbuscular Mycorrhizas

For 20 years the natural occurrence of AM fungi in soils highly polluted by heavy metals has been repeatedly reported (24–34). Considerable AM fungal species diversity can exist in metal-contaminated soils (30,32,35,36). In mine spoils containing elevated levels of Zn and Cd, Gildon and Tinker (24) found that the indigenous clover had high AM colonization levels. In a calamine spoil covered with a natural vegetation of metal-adapted plant species, Pawlowska et al. (32) found extremely high colonization levels in some plant species as well as spores of six different AM fungi. Different species of Glomus were also found in the rhizosphere of the zinc violet, Viola calaminaria, a metallophyte plant that colonizes metalliferous soils and ancient mining sites (33,34). Zak et al. (25) found that the number of spores of AM fungi in mine soils in Canada was very heterogeneous and probably leads to differences in the genetic pool of AM fungi. By using Medicago sativa as host plant, a mycorrhizal infection potential in mine spoils was demonstrated (37). The beneficial effect of mycorrhizal symbiosis on Zn tolerance of the grass Andropogon gerardii was proposed by Shetty et al. (36). The authors also showed that the AM fungi isolated from Zn-contaminated

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soil surrounding mine tailing were more effective in increasing the plant biomass at high levels of Zn, whereas AM fungi from a noncontaminated site promoted plant growth only in the soil containing lower Zn concentrations. In soils severely polluted by atmospheric deposition of metals from smelters, an abundant mycorrhizal colonization has often been found in metal-adapted plants (27,29,38,39). Griffioen et al. (38) highlighted the fact that the yield of AM fungal infection of Agrostis capillaris was comparable for the plants grown on uncontaminated soil and on a Zn/Cd-polluted soil. However, a negative correlation between infection rate and Cu concentration in soil was observed. In agricultural soils polluted by sewage sludge amendments or by emissions from smelters, high levels of mycorrhizal colonization are also observed (40,41). Some studies revealed that in these polluted fields the mycorrhizal colonization of crops could be reduced or delayed (28,42,43), whereas in other studies the addition of metal containing sludge did not affect in situ AM development (44). Therefore, it has been hypothesized that AM fungal ecotypes can exhibit varying degrees of tolerance to metals (43). The effects of different inputs of metal-contaminated sewage sludge to an agricultural field experiment (Braunschweig, Germany) have been studied. For plots that received the highest sludge amendment, the total number of spores in soil decreased significantly. However, metal concentrations in these soils remained below the European Community’s maximum allowable concentration limits for metals in soils (45,46). As pointed out by Del Val et al. (43), the inhibition of mycorrhizal colonization in the contaminated soils could be due in part to the inhibition of fungal spread. In addition, modification of the soil pH, P concentration, and CEC can be correlated with the mycorrhizal colonization (41,42,47). Leyval et al. (11) emphasized that the mycorrhizal infectivity of different heavy metal–polluted soils is linked to the amounts of bioavailable/extractable metals from soils rather than total metal concentration. Although this idea is now accepted, the availability of metals has been rarely measured, so that comparisons between studies are difficult to make (11). 3.2

Ectomycorrhizas

Ectomycorrhizal fungi can be found everywhere potential host trees can colonize metal-contaminated soils. However, relatively few tree species can survive on metalliferous soils with high metal loads. Metallophyte vegetations on the naturally metalliferous soils in western and northern Europe are slowly invaded by ectomycorrhizal Betula sp., Salix sp., Pinus sp. (32,48,49), and by arbuscular mycorrhizal Acer sp. (50). Related taxa can be found on mining sites in America and in soils severely polluted by atmospheric deposition of metals from smelters. In most cases, tree growth is seriously affected by metal stress, and metals in leaves or needles can reach toxic concentrations. In 1975, a field site was designed as a test plot for the reforestation of the ‘‘zinc desert’’ in Lommel (Belgium).

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The site is polluted with nonferrous metals emitted by a zinc smelter, which was dismantled in 1973 (39). Container plants of Pinus sylvestris, P. nigra, Betula pendula, Robinia pseudoacacia, Prunus serotina, and several Acer spp. were planted in homogeneous blocks of 20 ⫻ 20 m on a 1-ha plot, devoid of vegetation due to Zn toxicity. Saplings were planted in large planting holes (⫾10 L) filled with the original sandy soil amended with a metal immobilizing substance (loess, marl, or compost). Although there are a few miserable survivors of most tree species after 25 years, only the ectomycorrhizal genera Pinus and Betula were able to form a real ‘‘forest.’’ In 1985, we found basidiocarps of 7 ECM fungal species (6 genera) in these plots. In 1999, species richness had increased to 17 ECM species from 11 genera. The most abundant sporocarps in both years were from Suillus luteus and Amanita muscaria. A considerable number of reports mention the presence of basidiocarps of several ECM species on severely metalpolluted soils, but systematic field studies are scarce. In addition, there is clearly a need for more below-ground ECM community studies in these highly contaminated environments. Vra˚lstad et al. (51) found that a dominant ECM morphotype on seedlings of Pinus sylvestris, Betula pubescens, and Salix phylicifolia growing at the edge of a copper mine spoil in Norway was formed by an ascomycete from the Hymenoscyphus ericae aggregate. This mycobiont probably never forms sporocarps, and this finding suggests that we only have a very incomplete view on the biodiversity of mycorrhizal symbionts in metal-polluted environments. Screening experiments in vitro and in symbiosis have confirmed considerable variation in heavy metal sensitivity among ECM fungi (22,52–56). 3.3

Ericoid Mycorrhizas

The dwarf shrub Calluna vulgaris is a characteristic ‘‘ericoid’’ host plant that is a successful colonist of some acidic Cu-, Zn-, and/or Pb-polluted soils in Europe (12,48,51,57). On these soils, it appears that the plants are consistently colonized by a mycorrhizal symbiont, Hymenoscyphus ericae (Pezizella ericae), which clearly is essential for the survival of the plants in these contaminated habitats (57). Another ericoid mycorrhizal fungus from metal contaminated soils is Oidiodendron maius, which was present in roots of Vaccinium myrtillus (58). 4

HEAVY METAL TOLERANCE IN MYCORRHIZAL FUNGI

The in situ occurrence of AM fungal spores and/or mycorrhizal colonization even in highly polluted soils reported in numerous papers strongly suggests the possible metal tolerance of AM fungi. Gildon and Tinker (59) reported the first evidence of a possible tolerant state of an AM fungal isolate. These authors, along with others, have tested the growth performance of AM fungi by measuring different mycorrhizal parameters such as AM root colonization, and the ratio of

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coils, arbuscules, and vesicles within the root cortical cells for different AM fungal isolates and different host plants (24,43,59,60). Since AM fungi cannot complete their life cycle in axenic cultures (strict biotroph), metal tolerance has also been estimated from the ability of AM spores to germinate in a polluted substrate (61–63). Because the preinfective stages of the AM fungi are very sensitive to the presence of heavy metals (64), such an in vitro biotest could be very efficient to detect metal toxicity. By using one of these strategies, a higher tolerance to Zn, Cd, Cu, and/or Pb of indigenous fungi isolated from polluted soils in comparison to isolates from unpolluted soils has been clearly demonstrated in several

FIG. 2 Zn and Cd tolerance of isolates of the ectomycorrhizal fungus Suillus luteus. Tolerance indices (%) calculated for in vitro biomass production (dry weight). Open symbols, isolates from a metal polluted site (Lommel, B); closed symbols, isolates from a nonpolluted site (Paal, B).

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studies (59–61). Elsewhere different susceptibility to heavy metals was highlighted among isolates from the same metal-polluted soil (46). By using a Cdtolerant Glomus mosseae, Joner and Leyval (65) have shown that the growth of extraradical hyphae was not affected by high concentration of Cd added to the soil of the hyphal compartment. Nevertheless, to date there is no published evidence of the stability of metal tolerance in AM fungi. Metal tolerance in ECM fungi has been studied repeatedly (21,22). From these studies it appears that ECM fungi from metal-contaminated soils often display constitutive tolerance, although it cannot be excluded that some species only colonize less polluted microsites in the polluted soil. In most cases, there is little evidence for intraspecific adaptive changes between ECM populations from polluted or nonpolluted habitats (22). Nevertheless, adaptive tolerance to Zn and Cd was observed in some ECM fungi at the Zn- and Cd-polluted Lommel site in Belgium (53). These results were recently confirmed on the population level of Suillus luteus, and it also appears that the trait is genetically stable (66). In Fig. 2, the in vitro Zn and Cd tolerance is shown for S. luteus isolates from a polluted and nonpolluted site. Adaptive Al tolerance was also found in the same mycobiont (67) and in another pioneer ECM fungus: Pisolithus tinctorius (68). Adaptive Zn tolerance has also been observed in mycorrhizal endophytes of the Ericaceae from polluted soils (58). 5

PLANT PROTECTION TO METALS: FUNCTIONS OF MYCORRHIZAL FUNGI AND MECHANISMS INVOLVED

There has been increasing interest in the functions of AM fungi in heavy metal– polluted soils. There is circumstantial evidence that plants can benefit from AM colonization on highly contaminated soils. Nevertheless, it remains a matter of debate whether AM fungi protect their host from exposure to metal contaminants by reduced assimilation or transfer of metals to the host (12). AM colonization is maintained on mine plant vegetations for the same reasons that fungi are maintained in uncontaminated environments (e.g., phosphorus acquisition). Coevolution of hosts and AM fungi in metal-contaminated environments can nevertheless be expected and this process might lead to a further increase of the metal tolerance of a host. Using a compartmentalized system separating roots and extraradical mycelium, it has been demonstrated that the extraradical mycelium of AM fungi can transport 65Zn (69) and 109Cd (65) from soil to roots. Nevertheless, the metal transfer from roots to shoots was limited (65). The authors hypothesized that Cd was immobilized within the roots. To estimate the capacity of hyphae to bind metals, Joner et al. (70) performed an experiment with excised mycelium of different Glomus sp. isolates with different history of exposure to Zn and Cd. They showed that AM hyphae have a high metal sorption capacity (up to 0.5 mg of

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Cd per mg of dry biomass) and that the metal sorption capacity was related to the adaptation of the fungus to high levels of heavy metals. Other authors have studied the sorption and subcellular localization of Cu and As in the extraradical mycelium of different isolates of Glomus sp. by transmission electron microscopy combined with X-ray microanalysis (EDAX) (71). The results showed that both Cu and As were sequestered within the hyphal walls. In agreement with Joner et al. (70), Gonzalez-Chavez et al. (71) showed differences in the levels of sorbed Cu highlighting functional differences among AM fungi that can be related to the history of the exposure of the isolates. Tolerant AM fungi isolated from polluted soils sorbed more metals than nontolerant fungi, resulting in a limitation of metal transfer to plants (70,71). In 1994, Medeiros et al. (72) reported differences in efficiency of mycorrhizal fungi in the uptake of metals by plants. In addition, Shetty et al. (30) showed that mycorrhizal grasses were more successful than nonmycorrhizal grasses in the colonization of mine soil. The main ecological implication of these investigations was that the obvious plant metal tolerance might be conferred by the fungal symbiont. Recent studies have enforced this hypothesis (33,34,73). Metal-tolerant fungi allow the establishment of plants—of even nonadapted species—in highly metal-polluted soil. According to Wilkinson and Dickinson (49), there are more opportunities for genetic changes that match environmental constraints in fungi then in short- or long-lived higher plants. Therefore, one of the major functions of mycorrhiza would be to allow the adjustment to local soil conditions and acclimation of plants. Coevolution of ectomycorrhizal fungi and their long-lived tree hosts for more metal tolerance in polluted environments is probably a very slow process. There is little evidence that on metalliferous soils the tree species have adapted genetically to a similar extent as some herbaceous plant species did in the same habitats (49,74,75). The occurrence of Zn and Pb tolerant ecotypes of birch on metalliferous soils has been reported by Brown and Wilkins (76) and by Eltrop et al. (48). ECM fungi may adapt more rapidly than trees due to their shorter life cycles. Selection of more resistant ECM fungi is likely to be involved in the survival of highly mycotrophic trees such as pines and birches in metal-polluted environments. As with the AM symbiosis, there is evidence that ECM fungi have beneficial effects in the alleviation of metal toxicity in forest tree seedlings (13). Some authors argue that ECM fungi do not confer increased metal resistance to their host but that the symbiosis fulfills its normal ecological function (12). We believe that this conclusion cannot be generalized since particular ECM fungi are able to restrict transfer of metals to the host while maintaining normal nutrient transfer to their host tree (54,77,78). Metal transfer to shoots is, however, a complex process that is partly affected by the transpiration stream (13). However, as long as the mycorrhizal root system remains functional, a reduced transfer of metals to the host might promote survival of trees in polluted environments (79).

46

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More convincing evidence that ECM fungi can endow their hosts with metal resistance was recently demonstrated in a dose-response experiment in which mycorrhizal and nonmycorrhizal pine seedlings were exposed to four Cu concentrations. Plants were grown in semihydroponics with P as the growth-limiting element. Biomass production, P content, and P uptake rate were severely inhibited with increasing Cu concentrations in nonmycorrhizal (NM) seedlings, whereas plants colonized with Thelephora terrestris or Suillus bovinus were respectively not at all or much less affected by the Cu treatments. Cu transfer to shoots remained lower in mycorrhizal plants than in NM seedlings. Results of this experiment are illustrated in Fig. 3 showing the Cu response on the ratio of ECM/NM plants for biomass, P content, and Cu burden in the needles. In the absence of

FIG. 3 Response of Pinus sylvestris seedlings to toxic Cu concentrations. Ratio of ectomycorrhizal (ECM)/nonmycorrhizal (NM) plants for total biomass (■), total P (䉱) and total Cu in shoots (䊉). Seedlings were grown under P limitation and were exposed to elevated Cu for 36 days. (A) Plants mycorrhizal with Thelephora terrestris. (B) Plants mycorrhizal with Suillus bovinus.

Mycorrhizal Fungi

47

elevated Cu, the ECM/NM ratio for biomass and P content is close to 1, indicating no difference between NM and ECM plants. With increasing Cu exposure, the ECM/NM ratio for biomass and P content consistently increased, whereas total Cu in the shoots was relatively lower in ECM plants. These results showed that both mycobionts could protect root development and functioning of their host. At the highest Cu concentration (47 µM), transfer of Cu (and other nutrients) was impaired in NM plants because of a complete inhibition of root growth. Mycorrhizal colonization of the short roots was hardly affected by the Cu treatment. The mechanisms involved in the amelioration of metal stress remain unclear and are probably diverse. For a good overview of possible mechanisms, the reader is referred to the reviews of Leyval et al. (11) and Jentschke and Godbold (13). From our own experience, we suggest that ECM fungi that are able to survive at elevated metal exposure and that can avoid excessive uptake of metal in the symplast theoretically can reduce metal exposure of the roots of their host. In this respect it should be stressed that sporocarps of ECM fungi from metal-polluted sites generally do not accumulate the toxic metal, certainly when it is an essential element (66,80,81). Low metal concentrations or a low mobility of metals in mycelia might be important to reduce metal exposure to host roots. Prevention of excess metal uptake might be realized by excretion of metalimmobilizing substances, extracellular sequestration, or well-regulated uptake systems coupled to stable plasmalemmas. 6

DIVERSITY OF MYCORRHIZAL FUNGI IN HEAVY METAL POLLUTED SOIL

Because mycorrhizal fungi constitute a living bridge for the plant nutrition and resistance to different environmental stresses, there is an increasing interest in the diversity of mycorrhizal fungi in soils. The species richness of AM fungi has been analyzed for both nonpolluted soil (82–87) and heavy metal–polluted soil (30,32,34,36,45,46,88,89,90) (Table 1). Species richness is probably dependent on the type of soil, but it nevertheless seems to decrease when the soils are polluted (Table 1). In the few studies that focused on old, nondisturbed mine vegetations, spores of several AM fungi were found (32,34). At present, it is not clear whether some of these fungi have a preference for such habitats. In agricultural soils, the AM fungal species richness is usually low (45,46,86) compared with the number of species found in natural grasslands (82,83). Soil tillage and agricultural inputs of pesticides and fertilizers are suspected to deplete AM fungal diversity (91,92). In most cases, the species richness measured is the result of cumulated local diversity(90). In two different field experiments, a long-term sewage sludge application with increasing concentrations of metals lead to a decrease in both size and diversity of AM fungi populations (45,46). For soils receiving intermediate rates of

TABLE 1 Analyses of the AM Fungal Species Richness in Different Ecosystems

2

48

Location, surface (if known), sampling strategy, reference

Total no. of species

AM genera (%) G

A

E

Gi

Sc

12 21 43 14 26 25

58 67 63 57 38 44

25 5 9 14 8 24

0 5 0 14 4 4

0 14 7 0 11 16

17 9 21 14 38 12

Acidic mine soil, West Virginia (60 m2), S 1 sample, TC (88) Dunes, Joaquina, Brazil (600 m2), S 1 sample (89) Industrial soil, North Carolina (75 m2), S 1 sample, TC (90) Mine spoil, Kansas (⬍1 ha), S 1 sample (30) Calamine spoil, metal tolerant vegetation, Poland (32) Disturbed site, S 3 samples Undisturbed site, S 3 samples Mine spoil, Kansas (⬍1 ha), S 1 sample (36) Control soil, S 1 sample (36) Sewage sludge amended field experiment (old arable field trial), Braunschweig, Germany (85.5 m2) (45) Control plot, S1 sample, TC Moderately polluted soil, S1 sample, TC Highly polluted soil, S1 sample, TC Sewage sludge amended field experiment (ex-woodland field trial), Braunschweig, Germany (85.5 m2) (46) Control plot, S1 sample, TC Moderately polluted soil, S1 sample, TC Highly polluted soil, S1 sample, TC Mine spoil, rhizosphere zinc violet (Viola calaminaria) S 1 sample, TC (34)

5 12 24 3

20 33 42 100

60 17 29 0

0 0 0 0

0 8 12 0

20 42 17 0

2 6 3 5

50 83 67 60

0 0 0 0

50 17 33 40

0 0 0 0

0 0 0 0

6 6 5

100 100 100

0 0 0

0 0 0

0 0 0

0 0 0

3 4 2 4

100 100 100 100

0 0 0 0

0 0 0 0

0 0 0 0

0 0 0 0

a Studies performed in nonpolluted soils are reported on the top of the table. Taxonomic analysis performed directly on the sampled soil (S) or on trap cultures (TC). Ratio of the number of species that belong to Glomus (G), Acaulospora (A), Entrophospora (E), Gigaspora (Gi), and Scutellospora (Sc).

Colpaert and Vandenkoornhuyse

Prairie, Central Iowa (14336 m ), S 32 samples (82) ‘‘Konza Prairie,’’ Kansas (35 km2), S 6 samples (83) Apple plantings, 18 states of the USA, S 18 samples and TC (84) Dunes, ‘‘Great Lake,’’ Wisconsin (600 km), S 10 samples (85) Dunes, East coasts of the USA (355 km), S 19 samples (86) ‘‘Cedar Creek Natural History Area,’’ Minnesota (18 km2), S 15 samples (87)

Mycorrhizal Fungi

49

sludge contamination, Del Val et al. (46) found a decrease of the total spore number whereas the diversity index increased. On the other hand, Vandenkoornhuyse (45) highlighted an increase of the AM fungal population size and constant species richness in moderately polluted soil. A similar response to stress was revealed on the same field experiment for Rhizobium leguminosarum bv. trifolii (93). Additional analyses on the diversity of spores isolated from the soil of the different plots of the field experiment were performed. The analysis of the intraspecific diversity of two indigenous Glomus species by ISSR fingerprints on single spores revealed a high diversity whatever the AM fungal population (94). The number of haplotypes was correlated with the size of the population. For both species, the population genetic analyses showed a high likelihood of occurrence of recombination events for the populations of spores isolated from the most polluted soil. This mechanism might allow the fungi to generate more genetic diversity so that they can adapt more rapidly to the environmental stress (94). A hypothetical model of the effects of stress on an AM fungal population can be proposed. In this model, we assume that in stable and uniform environments the competition might allow the ‘‘neutral’’ emergence of competitive species or genotypes with high fitness. A moderate stress may modify the previous equilibrium by a possible decrease of the fitness of the individuals leading to a proliferation of more types (15,18). For a higher level of stress, a progressive extinction of nonadapted organisms ends in the loss of diversity. The number of ECM fungi that can associate with ectomycorrhizal trees is considerably larger than in the AM symbiosis. However, as pointed out above, we have a very incomplete view on the biodiversity of mycorrhizal symbionts in metal-polluted environments. Nevertheless, in expectation of more below-ground community studies, we would like to refer to two community studies based on above-ground sporocarp observations, along metal gradients in northern and southern Sweden (81,95). Although molecular studies have demonstrated a considerable lack of correspondence between the above- and below-ground communities of ECM colonizers (16), the field studies of Ru¨hling suggest that particular mycorrhizal guilds disappear with increasing metal stress. Sporocarp production in Picea abies forests along a Cu-Zn gradient was investigated (95). The average number of species of macrofungi per plot (1000 m2) decreased significantly along the gradient: 35 species at a Cu concentration of about 100 µg/g organic matter, 25 species at 1000 µg/g, and only 15 species at 10,000 µg Cu/g organic matter. Some ECM taxa decreasing along the gradient were Chantarellus cibarius, Cortinarius sp., Dermocybe sp., Gomphidius sp., Hydnum sp., Lactarius sp., Paxillus involutus, and Russula sp. Taxa that were not affected or that increased in frequency include Albatrellus ovinus, Amanita sp. Cantharellus tubaeformis, Laccaria laccata, and Leccinum sp. In northern Sweden, a similar study was performed

50

Colpaert and Vandenkoornhuyse

along a more complex metal (As, Cu, Pb, Cd, Zn) gradient and very similar results were obtained as well (81). In the least polluted plots, 4.4 species of macrofungi were found per 100 m2, whereas only 1.3 species were observed in the most heavily polluted plots. The number of observations of the genus Amanita increased in the most heavily polluted area, whereas Cortinarius, Lactarius, and Russula showed decreasing numbers of observations. The reduced sporocarp production and the decreasing above-ground diversity do not necessarily mean that the percentage of root colonization decreases over the same gradient. It is certainly possible that the frequency of other nonfruiting fungi (Hymenoscyphus ericae) increases on the most polluted sites, as was observed in the study of Vra˚lstad et al. (51). Other ectomycorrhizal taxa that have been frequently found on heavily polluted soils include Hebeloma sp., Pisolithus tinctorius (96), Rhizopogon sp. (97), Scleroderma sp. (98), Suillus sp. (22,66) and the Cd-accumulating Amanita muscaria (53,80).

7

AM FUNGI AND SOIL QUALITY ASSESSMENT

The different ways by which AM fungi influence their host plants and their terrestrial ecosystems illustrates fairly well the importance of these fungi for soil quality. However, little attention has been paid on this group of microorganisms in the assessment of soil quality. In agricultural soils, different practices may harm the potential AM colonization (and consequently functions) (92). Phosphate fertilization (99), soil pH, availability of carbon and nitrogen (87), soil tillage (91), crop rotation (100), and pesticide application (101) all can modify or reduce the AM formation. Among current agricultural practices, the application of sewage sludge and associated heavy metal pollution is becoming an important problem. Since numerous factors can modify the AM fungi in soils, the measurement of a toxic effect on AM fungi is difficult to assess properly. In a field experiment with different plots of the same soil structure and the same crop history, it is possible to study the long-term effect of metal-contaminated sewage sludge applications on the indigenous AM fungi. In Table 2 we show the metal concentrations measured in an old arable field experiment (45) and in an ex-woodland field experiment of Braunschweig (Germany) (46). In both studies, the metal concentration in the most polluted plot was below the European Community’s maximum allowable concentration limits for metals in sludge-treated soils (Table 2). As described before, a toxic effect was nevertheless observed on the populations of AM fungi in these soils (45,46,94). Metal toxicity is even suspected in moderately polluted soils. The measurement of fluxes of free metals in soil solution (102) demonstrated that Zn and Cd are the most available elements. However, it is difficult to determine which factors are responsible for the toxicity. We and others (18,103) argued that decisions on soil protection should be based on knowledge

Available heavy metals pFss (pFss ⫽ ⫺log Fss)

Total metal conc. (mg/kg)

Old arable soil: C MP HP Old woodland soil: C MP HP Regulations: European communityb United Statesc

Zn

Cd

Ni

Cu

Pb

Zn

Cd

Ni

Cu

Pb

56 163 329

0.4 0.8 2.7

8.5 16.3 32

11.9 42.9 93.7

22 29 95

3.3 2.9 1.2

5.0 4.9 3.4

3.8 3.6 2.1

3.6 3.4 2.5

4.2 4.1 3.6

43 88 295

0.3 0.6 2.8

8.6 8.3 29

9.8 17 92

29 30 111

2.4 1.9 1.1

3.9 3.8 3.0

3.5 3.1 1.8

3.6 3.3 2.5

3.8 3.7 3.4

150–300 1400

1–3 20

30–75 210

50–140 750

50–300 150

— —

— —

— —

— —

— —

Mycorrhizal Fungi

TABLE 2 Concentrations of Metals in an Agricultural Soil Treated with Sewage Sludge (Braunschweig, Germany)a

a Heavy metals in a control plot (C), a moderately polluted plot (MP), and a highly polluted plot (HP) in old woodland soil (43,46) and in old arable soil (45,94). The flux of metal from solid soil to solution (Fss) in ng cm⫺2s⫺1 estimates the available heavy metal fraction (102). Maximal concentrations of heavy metals allowed in agricultural soils are included for comparison. b Commission of the European Communities. Council directive (86/278/EEC) on the protection of the environment, and in particular of the soil, when sludge is used in agriculture. Off J Eur Comm L181(annex 1A) pp 6–12, 1986. c Calculated from maximum cumulative pollutant loading limits without taking into account background concentrations of the elements in soils. U.S. environmental protection agency. Standards for the use or disposal of sewage sludge. Fed Regist 58:9248–9415, 1993.

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68. 69.

70. 71.

72. 73.

74. 75. 76. 77.

78.

79.

80. 81.

82. 83.

84.

Colpaert and Vandenkoornhuyse in Suillus luteus (L.) S.F. Gray., an ectomycorrhizal symbiont of Scots pine. Acta Soc Bot Pol 64:97–105, 1995. LM Egerton-Warburton, BJ Griffin. Differential responses of Pisolithus tinctorius isolates to aluminum in vitro. Can J Bot 73:1229–1233, 1995. B Bu¨rkert, A Robson. 65Zn uptake in subterranean clover (Trifolium subterraneum L.) by three vesicular-arbuscular mycorrhizal fungi in a root-free sandy soil. Soil Biol Biochem 26:1117–1124, 1994. EJ Joner, R Briones, C Leyval. Metal binding capacity of arbuscular mycorrhizal mycelium. Plant Soil 226:227–234. C Gonzalez-Chavez, J D’Haen, J Vangronsveld, J Dodd. Copper adsorption and accumulation by external mycelium of three arbuscular mycorrhizal fungi from polluted soils. COST 838 Meeting, Arbuscular Mycorrhizas and Plant Health under Abiotic Stress. Centre de Pedologie Biologique, CNRS Nancy, 10–11 December 1999, Abstract book, p. 4. CAB Medeiros, RB Clark, JR Ellis. Effects of excess manganese on mineral uptake in mycorrhizal sorghum. J Plant Nutr 17:2203–2219, 1994. M Kaldorf, AJ Kuhn, WH Schro¨der, U Hildebrandt, H Bothe. Selective element deposits in maize colonized by a heavy metal tolerance conferring arbuscular mycorrhizal fungus. J Plant Physiol 154:718–728, 1999. E Simon. Heavy metals in soil, vegetation development and heavy metal tolerance in plant populations from metalliferous areas. New Phytol 81:175–188, 1978. NM Dickinson, AP Turner, NW Lepp. How do trees and other long-lived plants survive in polluted environments? Funct Ecol 5:5–11, 1991. MT Brown, DA Wilkins. Zinc tolerance in Betula. New Phytol 99:91–100, 1985. MD Jones, TC Hutchinson. Nickel toxicity in mycorrhizal birch seedlings infected with Lactarius rufus or Scleroderma flavidum. I. Effects on growth, photosynthesis, respiration and transpiration. New Phytol 108:451–459, 1988. MD Jones, TC Hutchinson. Nickel toxicity in mycorrhizal birch seedlings infected with Lactarius rufus or Scleroderma flavidum. II. Uptake of nickel, calcium, magnesium, phosphorus and iron. New Phytol 108:461–470, 1988. KK Van Tichelen, T Vanstraelen, Colpaert JV. Nutrient uptake by intact mycorrhizal Pinus sylvestris L. seedlings: a diagnostic tool to detect copper toxicity. Tree Physiol 19:189–196, 1999. CH Gast, E Jansen, J Bierling, L Haanstra. Heavy metals in mushrooms and their relationship with soil characteristics. Chemosphere 17:789–799, 1988. A Ru¨hling, B So¨derstro¨m. Changes in fruitbody production of mycorrhizal and litter decomposing macromycetes in heavy metal polluted coniferous forests in north Sweden. Water Air Soil Pollut 49:375–387, 1990. C Walker, CW Mize, HS McNabb. Populations of endogonaceous fungi at two locations in central Iowa. Can J Bot 60:2518–2529, 1982. BAD Hetrick, J Bloom. Vesicular-arbuscular mycorrizal fungi associated with native tall grass prairie and cultivated winter wheat. Can J Bot 61:2140–2146, 1983. DD Miller, PA Domoto, C Walker. Mycorrhizal fungi at eighteen apple rootstock plantings in the United States. New Phytol 100:379–391, 1985.

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85. RE Koske, LL Tews. Vesicular-arbuscular mycorrhizal fungi of Wisconsin sandy soils. Mycologia 79:901–905, 1987. 86. RE Koske. Distribution of VA mycorrhizal fungi along a latitudinal temperature gradient. Mycologia 79:55–68, 1987. 87. NC Johnson, DR Zak, D Tilman, FL Pfleger. Dynamics of vesicular-arbuscular mycorrhizae during old field succession. Oecologia 86:349–358, 1991. 88. JB Morton. Three new species of Acaulospora (Endogonaceae) from high aluminum, low pH soils in West Virginia. Mycologia 78:641–648, 1986. 89. SL Stu¨rmer, MM Bellei. Composition and seasonal variation of spore populations of arbuscular mycorrhizal fungi in dune soils on the island of Santa Catarina, Brazil. Can J Bot 72:359–363, 1994. 90. JB Morton, SP Bentivenga, JD Bever. Discovery, measurement, and interpretation of diversity in arbuscular endomycorrhizal fungi (Glomales, Zygomycetes). Can J Bot 73 (Suppl. 1):S25–S32, 1995. 91. NC Johnson, FL Pfleger. Vesicular-arbuscular mycorrhizae and cultural stresses. In: GJ Bethlenfalvay, RG Linderman, eds. Mycorrhizae in Sustainable Agriculture. Special publication No. 54 American Society of Agronomy, Madison, Wisconsin, 1992, pp. 71–99. 92. M Kling, I Jakobsen. Arbuscular mycorrhiza in soil quality assessment. Ambio 27: 29–34, 1998. 93. AM Chaudri, SP McGrath, KE Giller, E Rietz, DR Sauerbeck. Enumeration of indigenous Rhizobium leguminosarum biovar trifolii in soils previously treated with metal-contaminated sewage sludge. Soil Biol Biochem 25:301–309, 1993. 94. P Vandenkoornhuyse, C Leyval, I Bonnin. High genetic diversity in arbuscular mycorrhizal fungi: evidence for recombination events. Heredity (submitted). 95. A Ru¨hling, E Ba˚a˚th, A Nordgren, So¨derstro¨m B. Fungi in metal-contaminated soil near the Gusum brass mill, Sweden. Ambio 13:34–36, 1984. 96. K Turnau, E Gucwa, P Mleczko, B Godzik. Metal content in fruit-bodies and mycorrhizas of Pisolithus arrhizus from zinc wastes in Poland. Acta Mycol 33:59– 67, 1988. 97. K Turnau, J Faber, J Dexheimer, B Botton. Influence of Zn and Pb on Rhizopogon roscolus mycelium—energy dispersion spectroscopy and cytochemical investigation. Acta Soc Bot Pol 68:57–62, 1994. 98. MD Jones, Hutchinson TC. The effect of mycorrhizal infection on the response of Betula papyrifera to nickel and copper. New Phytol 102:429–442, 1986. 99. I Jakobsen. Vesicular-arbuscular mycorrhiza in field grown crops. III. Mycorrhizal infection and rates of phosphorus inflow in pea plants. New Phytol 104:573–581, 1986. 100. JA Ocampo, DS Hayman. Influence of plant interactions on vesicular-arbuscular mycorrhizal infections. II. Crop rotations and residual effects of non-host plants. New Phytol 84:27–35, 1980. 101. JM Trappe, R Molina, M Castellano. Reactions of mycorrhizal fungi and mycorrhiza formation to pesticides. Annu Rev Phytopathol 22:331–359, 1984. 102. H Zhang, W Davison, BP Knight, SP McGrath. In situ measurement of solution concentrations and fluxes of trace metals in soils using DGT. Environ Sci Technol 32:704–710, 1998.

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103. E Witter, KE Giller, SP McGrath. Long term effects of metal contamination on soil microorganisms. Soil Biol Biochem 26:421–422, 1994. 104. E Witter. Towards zero accumulation of heavy metals in soils: an imperative or a fad? Fertil Res 43:225–233, 1996. 105. B Kendrick. The Fifth Kingdom. 2nd ed. Newburyport, MA: Focus Information Group, 1992, p. 406.

3 Freshwater Algae Barbara Pawlik-Skowron´ska and Tadeusz Skowron´ski Institute of Ecology of the Polish Academy of Sciences, Lublin, Poland

1

INTRODUCTION

Algae are phototrophic organisms that are very diverse in size and morphology. Most of them are aquatic and inhabit both fresh and saline waters. They are classified in the following divisions: Chlorophyta, Chromophyta, Rhodophyta. Depending on the classification, prokaryotic blue–green forms (Cyanobacteria, Cyanophyta) and Prochlorophyta may be also included. Some metals are required for algal metabolism and physiology as macroelements (K, Mg) or microelements (Co, Cu, Fe, Mn, Mo, Ni, V, Zn), and must be obtained from the external environment. However, the essential elements, such as Cu, Mn, Ni, and Zn, can be toxic at high concentrations. Other metals, such as Cd, Hb, Pb, have no known biological function and are always toxic. The term ‘‘heavy metals,’’ used by ecologists, generally refers to the elements possessing a density greater than 6 g/cm3 and connotations of toxicity, e.g., Ag, As, Cd, Cu, Hg, Mn, Ni, Pb, and Zn (1). As reported in several reviews (2–8), all of these and related metals as well as metalloids may be accumulated in algae and have an impact on them. 59

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ACCUMULATION OF METALS BY ALGAE

Algae, primary producers and essential constituents of setting materials in aquatic ecosystems, are often the dominant vectors for heavy metals. Plankton particles (including cyanobacteria and eukaryotic algae) play an important role as biosorbents influencing the environmental fate of metals, their chemical speciation and bioavailability (9–11). For example, during a plankton abundance maximum sorption of cadmium occurred, with phytoplankton playing a significant role. Both nanoplankton and phytoplankton were largely responsible for transferring cadmium from the water column to the sediments (12). On the other hand, algae as primary producers can introduce heavy metals into the aquatic food chain and this phenomenon may have important consequences for ecosystem structure and function (13). Algae possess the capacity for taking up heavy metals from the water, producing an internal concentration greater than that of their surroundings. For example, the samples of Cladophora glomerata collected from different rivers and streams in northeastern England contained 1–33 µg Cu/g; 6.63–1130 µg Zn/g; 0.042–6.97 µg Cd/g; and 0.141–330 µg Pb/g; whereas the metal concentrations in the filterable water were 3.1–17.9, 2–167; 0.04–8.3, and 0.5–22.4 µg/L, respectively. Significant correlation between the concentration of metal in Cladophora and water was stated (14); therefore, the algal ability for heavy metal accumulation may be useful for environmental monitoring, and chemical analyses of these organisms may give valuable information about contamination of the aquatic environment (1). Metal concentrations in algal biomass can vary widely. For instance, Zn concentrations in different taxons of algae and cyanobacteria at different locations ranged from 5.8 to 219,000 µg/g (4); Cd concentrations from 1 to 340 µg/g (5); and Pb concentrations from 1 to 14,000 µg/g (6). As recently reported (15), the cellular Cu concentration in the green alga Scenedesmus subspicatus can range from 2 to 950 µg/g at the external free Cu concentrations 10⫺14 –10⫺7 M, and from 33 to 6500 µg Zn/g at the external free Zn concentrations 10⫺9 –10⫺5 M. The metal bioaccumulation ability of algae in the environment can reflect to some extent a concentration factor (CF), i.e., the ratio between metal concentration in the cells and that in the surrounding water. This factor varied for various species and for different metals in a broad range (Table 1). CF is highly dependent on many environmental variables affecting the metal sorption level (free metal concentration in the solution, physiological status of the cell, algal cell density, kind of water, interference with other ions, etc.). Therefore, it is difficult to compare the CF values directly. The heavy metal uptake capacities of different algal species can be better described by sorption isotherms (16–19) showing a relationship between the metal concentration in the biomass and the residual metal concentration in water.

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TABLE 1 Studies Showing the Variety of Concentration Factors of Heavy Metals in Freshwater Algae Alga Chlorella pyrenoidosa Chlorella vulgaris Cladophora Nostoc 586 Scenedesmus obliquus Hydrodictyon reticulatum Tribonema Cladophora Hydrodictyon reticulatum Tribonema Hydrodictyon reticulatum Chlorella pyrenoidosa Nostoc 586 Hydrodictyon reticulatum Anacystis nidulans Ulothrix Coccomyxa Nostoc 586 Tribonema

Metal

Concentration factors

Ref.

Cd Cd Cd Cd Cd Cr Cu Cu Cu Fe Fe Hg Hg Mn Ni Ni Pb Pb Zn

16,700–32,000 86–3900 500 92.5 105 132–425 13,000–83,000 1800–3500 498–3300 140,000–290,000 267–6949 24,000–47,300 318 3715–4945 1950–56,130 140 31,000 91.4 520–1100

57 16 58 57 57 59 60 60 59 60 59 57 57 59 61 60 60 57 60

The mechanism of heavy metal uptake by algal cells is complex process comprising two phases: a fast, metabolic independent surface binding of physicochemical nature and a metabolism-dependent phase, whereby the metal ions are transported across the cell membrane to the cells (3,20–27). From the quantitative point of view, in the case of planktonic green and blue–green algae the surface sorption may be the largest proportion of the total metal uptake (about 80–98%), as reported for Cd, Cu, Zn, and Pb (16,23,27,28). However, in the case of very low concentrations of essential metals (15), up to 80% of the total cellular copper in Scenedesmus subspicatus was located intracellularly and only 20% was located on the cell surface. With the increasing external free Zn concentrations from 10⫺9 to 10⫺5 M, the amount of superficially adsorbed Zn increased from 5% to 80% of the total cellular Zn content in this alga. The large surface area of microalgae may highly affect the metal equilibrium concentration in the surrounding water and, in consequence, the metal availability to other organisms. This concerns both living and dead algal biomass; dead cells can often adsorb metal ions in larger quantities than living biomass

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(19,23) because more binding sites are available. Algal surfaces possess various functional groups, such as carboxyl, amino, phosphate, hydroxyl, and sulphydry, capable of binding heavy metals (29–31). In sorption process (where an ion exchange mechanism usually predominates), metals usually bind to anionic sites by displacing protons from acidic groups or existing metals from anionic sites (29,32). For example, this has been reported for freshwater filamentous green algae Vaucheria, Tribonema, Spirogyra, Oedogonium (33), as well as some unicellular green and blue–green algae (19). Some metals and metalloids (e.g., tin, technetium, germanium, arsenic, selenium) can exist in the form of anionic complexes. They can be also taken up by algae because certain functional groups; such as amines and imidazoles, can be positively charged when protonated and may electrostatically bind negatively charged metal complexes (31,34). Particular taxonomic algal groups and species exhibit considerable variations of the surface properties and hence different metal binding characteristics. The structural and chemical variations of cell surface components as well as various cell morphologies may be responsible for different metal sorption capacities observed. For example, Eisenia bicyclis was more effective in binding aluminum at pH 2 than were Cyanidium caldarium, Spirulina platensis, and Chlorella pyrenoidosa, while Cyanidium caldarium bound more copper at pH 2 than did Eisenia bicyclis, Spirulina platensis, and Chlorella pyrenoidosa (31). There are marked variations in specific surface values (m2 cm⫺3) between individual algal species (35). For example, the cyanobacteria Anabaena inaequalis and Anabaena lutea displayed especially high Cd adsorption per surface unit, whereas some Chlorophyceae, i.e., Klebsormidium sp., Ulothrix gigas) had rather low surface adsorption capability. The cation binding capacities of the unicellular cyanobacteria Anacystis nidulans, Aphanocapsa sp., Synechocystis aquatilis, and the green algae Stichococcus bacillaris and Chlorella kessleri were much higher than those of the filamentous macroalga Vaucheria sp. (Xanthophyceae), which possess much lower ratio of the surface to volume (19,36). The metabolism-independent surface sorption of heavy metals reported for many eukaryotic and prokaryotic algae is rapid (3,16,17,19,22,23,28,37–39). Algae bind most metals from the solution within a few minutes and sorption equilibrium is soon reached. This passive phase of the metal uptake is usually described by sorption isotherms (the Langmuir or Freundlich model). Heavy metal sorption by algal cells may be influenced by many environmental variables, and pH is one of the most important factors because of its effect on the charge of functional groups and chemical speciation of metals. Low external pH decreased algal sorption of such metals as Cu, Zn, Cd, Pb, and Co (5,17,19,31). However, sorption of anionic metal complexes such as TcO⫺4 increased with the decreasing external pH (34). Temperature effect on the surface sorption of heavy metals is relatively small (17). However, other environmental factors, such as competing metal cations and complexing ligands, may strongly affect this process.

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The second, metabolic-dependent (active) phase of metal uptake occurs obviously only in the living algal cells. There have been demonstrated the active uptake of Mn2⫹ into Chlorella pyrenoidosa cells (40), energy-dependent Cd transport into S. bacillaris cells (24,41), Ni2⫹ transport into Anabaena cylindrica cells (42), Cd and Zn transport into Chlorella vulgaris cells (43), Cd transport into Synechocystis aquatilis cells (26), Zn active uptake by Selenastrum capricornutum (44) and Chara corralina (45), as well as Cu active transport into the cells of Nostoc calcicola (38). The intracellular transport of metal ions studied both in green and blue–green algae runs usually more slowly than surface sorption. In darkness, intracellular Cd, Zn, Cu, and Pb uptake was reduced as observed in both eukaryotic algae such as Chlorella, Stichococcus, Vaucheria (22–24,37), and prokaryotic ones (26,38). In both the blue–green algae Anabaena cylindrica (42), Nostoc calcicola (38), Synechocystis aquatilis (26), and the green algae Stichococcus bacillaris (24) and Chara corralina (45), membrane potential, proton gradient across the cell membrane, and ATP may be involved in intracellular uptake of heavy metals such as Cd, Cu, Ni, and Zn. The intracellular transport of divalent metal cations is concentration-dependent, frequently shows saturation with the increasing metal concentration, and follows the Michaelis-Menten kinetics. It was reported for Cu transport into Nostoc calcicola cells (38), Cd uptake into Anacystis nidulans (46) and Synechocystis aquatilis (26), Ni transport into Anabaena cylindrica (42), and Zn uptake into Selenastrum capricornutum (44). On the contrary to the surface sorption, metal intracellular transport into the algal cells is strongly temperature-dependent (26,41). Cd transport into the green alga Stichococcus bacillaris cells was completely inhibited at 4°C and increased with the temperature increase by 35°C (41). There is the assumption based on the competitive uptake studies between metal cations that nonessential metals (Cd, Pb) share in algal cells transport systems for essential elements like Ca, Mn, Zn (26,40,41,47,48). However, the process of essential metal uptake in algae is not fully recognized. It has been suggested (49) that the Ca influx into the cells of Chara corralina occurs via voltagedependent calcium channels; however, it is not clear if there are specific channels for Ca. Essential elements like Zn may be accumulated intracellularly into algal cells using high- or low-affinity transport systems (44,45). The pH of the surrounding environment can strongly influence not only the metal surface sorption (as mentioned earlier) but the metabolic-dependent intracellular transport as well because of its effect on metal speciation and cellular membrane potential (2). There are scarce data on this subject. As reported for the green alga Stichococcus bacillaris (41) and the blue–green alga (cyanobacterium) Synechocystis aquatilis (26), the intracellular Cd transport occurred at pH 7 but was strongly inhibited at the acidic pH values (pH 4–5) (Fig. 1). However, there was no effect of changing external pH on the Ca influx into the cells of Chara corralina (49).

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FIG. 1 Cd2⫹ transport (1) and 14C fixation (2) in the presence of cadmium in S. aquatilis at different external pHs. Total cadmium concentration (8.9 µM) and 14C fixation in the control without Cd at each pH values were taken as 100%. Transport measurements were made after 40 min. (From Ref. 91.)

Some metals and metalloids, such as Pb, Hg, Sn, and As, can occur as organometallic compounds, which can be also accumulated by algae and may exert even stronger toxic effects than inorganic forms of these elements (25,50,51,52). The accumulation of organotin compounds was reported for a few freshwater green algae like Ankistrodesmus falcatus (53), Scenedesmus obliquus, Chlorella vulgaris (54), and some cyanobacteria (55). The accumulation of organocomplexes of Pb, Cu, Cr, and Tl in neutral lipids and plastoglobuli of the green alga Cladophora has also been reported (50). However little is known about the uptake mechanisms of organometallic compounds. There are suggestions that lipophilic organometallic compounds (like methylmercury) can rapidly pass algal cell membranes via passive diffusion (56). 3

EFFECTS OF METALS ON ALGAL GROWTH AND METABOLISM

The primary impact of heavy metals on the algal cells is at the biochemical and physiological levels. Heavy metals can interact with cellular macromolecules or

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disturb cell membrane integrity, leading to inhibition of essential cellular metabolic or physiological process and ultrastructural changes. Uptake of heavy metals can result in inhibition of several enzymes (62). Generally, as mechanisms of metal action on enzymes, binding of the metal to the sulfhydryl group involved in the catalytic action or structural integrity of enzymes and deficiency of essential metals in metaloproteins or metal-protein complexes combined with substitution of the toxic metal for the deficient element, are proposed. Strong affinity to SH groups is exhibited particularly by such metals as Cd, Pb, and Hg (62). Besides, heavy metal uptake can disrupt cell transport processes; also, in some cases oxidative stress is possible (63). Free radicals generated by some transition metals can damage cell compartments leading to decrease in growth or cell death. In the presence of Cu, Co, Pb, and Ni, the increase of peroxidase activity in Selenastrum capricornutum to prevent cell damage was reported (64). For these reasons, heavy metals may interrupt the normal metabolic processes and integrity of algal cells in a number of ways. The uptake of toxic amounts of metals by algae can resulting among other things, inhibition of photosynthesis, as well as alteration of respiration and adenylate metabolism. In addition, several ultrastructural and morphological changes can be observed in the algal cells. Metal stress can affect total cell volume, number and relative volume of polyphosphate bodies, lipids, vacuoles, cell wall, and periplasmalemmal space, among other factors (65). Enlarged algal cells, unable to divide, have been often observed (66–69). The mentioned effects lead to the inhibition of algal growth. Some toxic effects of heavy metals on the algal cells are presented in Table 2. The toxic effects of heavy metals on algae depend on the condition of exposition, time, and heavy metal dose. Generally, the toxicity of most metals is a function of free metal ion concentration in the solution, rather than total metal concentration, and can vary in the presence of complexing agents (70,71). In the aquatic systems, a number of environmental factors affect metal toxicity through the influence on metal chemical speciation and metal uptake process. One of the most essential factors is pH. Generally, toxicity of such metals as Cd, Cu, and Zn toward algal cells decreased as the pH of surrounding solution decreased (2). For example, cadmium exerted toxic effects on the green alga Stichococcus bacillaris at pH 6–7, whereas it did not affect the biomass yield substantially at pH 3 and 9 (72). Cobalt, copper, and nickel were less toxic to the green alga Chlamydomonas reinhardti at pH 5 than at pH 7 (71). Also the decrease of copper and uranium toxicity at pH 5.7 compared to those at pH 6.5 was reported for the freshwater alga Chlorella sp. (73). Apart from hydrogen ions, other cations and anions can influence the toxic effect of metal on the aquatic organisms through various types of competitive interactions. These other ions (Ca2⫹, Mn2⫹) may protect the organism from metal uptake or from its toxic expression. For example, they markedly reduced cadmium toxicity to the green alga Stichococcus bacillaris (74). However, in some cases the presence of other ions

66

TABLE 2 Some Toxic Effects of Heavy Metals on Green and Blue–Green Algae Organism

Metal

Observed effect Growth inhibition inhibition; cell division inhibition inhibition inhibition

Cd Cd, Zn Cu Pb and organolead complexes

Growth Growth Growth Growth

Cu and Cu complexes Zn, Cd, Hg Cu Ni, Hg Zn Cu, Cd, Zn

Chlorella pyrenoidosa

Cd

Scenedesmus acutus Euglena gracilis

Cu Zn, Cd, Hg

Chlorella sp. Chlorella pyrenoidosa Chlorella pyrenoidosa

Hg, Zn Cu Pb

Chlorella sp. Scenedesmus acutus Chlorella sp.

Hg, Zn Cu Hg, Zn

Growth inhibition Cell division inhibition Cell division inhibition Growth inhibition Growth inhibition Growth inhibition Photosynthesis Inhibition of O2 evolution; inhibition of CO2 fixation Inhibition of O2 evolution Inhibition of O2 evolution; inhibition of CO2 fixation Inhibition of O2 evolution Inhibition of CO2 fixation Inhibition of chlorophyll biosynthesis and O2 evolution Destruction of chlorophyll Destruction of chlorophyll Inhibition of the rate of chlorophyll synthesis; reduction of thylakoid surface area

80 66,67,84 82 50 69 83 69 84 77 28 80 82 85 86 69 87 86 82 86

Pawlik-Skowron´ska and Skowron´ski

Chlorella pyrenoidosa Stichococcus bacillaris Scenedesmus acutus Chlorella fusca Cladophora glomerata Chlorella pyrenoidosa Euglena gracilis Chlorella pyrenoidosa Anabaena inaequalis Microcystis aeruginosa Chroococcus paris

Ref.

Cd Ni, Hg Cu, Cd, Zn

Anabaena flos-aquae Synechocystis aquatilis

Cd Cd

Chlorella fusca

mixture of metals: Al, Fe, Cu, Zn, Ni, Mn Pb and organolead complexes

Chlorella fusca Cladophora glomerata

Al

Plectonema boryanum Anabaena flos-aquae Anabaena variabilis Anabaena flos-aquae

Cd Zn Cd

Synechocystis aquatilis Chlorella sp. Euglena gracilis Chlorella pyrenoidosa

Cd Zn, Hg Zn, Cd, Hg Pb

Cu, Cd

Changes in subcellular structure and larger proportion of heterocysts Increased number of polyphosphate bodies Increased number of polyphosphate bodies Reduction in subcellular structures; cell wall lysis Other Metabolic Processes Changes in cellular adenylate metabolism Respiration inhibition Respiration inhibition Stimulation of dark respiration

88 84 89 90 75,91

92 50

93

68 94 95 90

26 86 85 87

67

Monoraphidium dybowski Stichococcus sp. Anabaena 7120

Reduction of thylakoid surface area Photosynthesis inhibition; pigment bleaching Inhibition of Hill activity and photosynthetic oxygen evolution Reduction in number of carboxysomes Reduced CO2 fixation, O2 evolution and carbonic anhydrase activity Ultrastructural Changes Chloroplast damage; disruption of thylakoidal membranes Breakdown of organelles; distendent endoplasmic reticulum; shrinkage of chloroplast; multinucleate cells; changes in cell wall structure Detachment of plasmalemma from the cell wall; destruction of cell wall; vacuolization

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Anabaena flos-aquae Anabaena inaequalis Anacystis nidulans

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may enhance the toxic effect. Among inorganic ligands phosphate and chloride ions are persistent complexing agents of heavy metals. They are important complexing factors also in freshwaters, particularly in eutrophic waters or these contaminated with mining effluents and sewages of high salinity. As reported for the blue–green alga Synechocystis aquatilis in solutions containing cadmium chloride complexes (CdCl⫹, CdCl2, CdCl3⫺), Cd toxicity was significantly limited (75). In the aquatic environment the dissolved organic matter has also been demonstrated to affect metal toxicity (76). Various chemical forms of metal differ in their availability to living organism (77). On the other hand, the algal sensitivity to metals is species- or strain-specific (78,79). The stress caused by heavy metals in algal cells induces quantitative and qualitative changes in the structure and functioning of algal communities. 4

DIVERSITY OF ALGAL COMMUNITIES IN THE FRESHWATER HABITATS POLLUTED WITH METALS

Although high concentrations of heavy metals are known to be toxic to biota, reports on the freshwater algae surviving in the environments with elevated concentrations of heavy metals are common (96–105). An important aspect of ecological investigations of algal communities in the metal-contaminated freshwaters is the descriptive studies of algal flora in the polluted rivers, streams, and lakes. Freshwater streams and rivers may carry heavy metals due to the leaching of natural outcrops or industrial (mining, smelting) and agricultural (artificial fertilizers, pesticides) activities. Most studies of algal communities in the metalcontaminated waters were carried out in the polluted streams or rivers situated in the Cu, Pb, and Zn mining regions (100,101,103,106–108). 4.1

Blue–Green Algae (Cyanobacteria)

In the metal-contaminated freshwater habitats of high pH, blue–green algae (cyanobacteria) are often abundant. The dominant genera reported in the zincenriched waters were filamentous nonheterocystous Oscillatoria, Phormidium, Plectonema, and Schizothrix (109,110). Blue–green algae were found in the habitats of very high zinc concentrations (10–100 mg Zn/L) but only of pH higher than 5 (111). Low pH (⬍5) severely limits their survival (112). Yasuno and Fukushima (113) showed that the blue–green algae Chamaesiphon minutus and Phormidium luridum usually appeared in the copper- and zinc-polluted rivers in Japan. As reported by Takamura et al. (103), different species of Phormidium occurred in three Japanese rivers polluted with Cu and Zn. Also in the water of the urban river Miyata of alkaline pH, polluted with Cu and Zn (108), the following cyanobacteria species were identified: Chamaesiphon subglobosus, Phormidium foveolarum, and P. uncinatum.

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Green Algae

In the river water in Wales (U.K.) polluted with Zn, Cd, Cu, and Pb, McLean and Jones (97) reported the occurrence of the filamentous green algae Hormidium rivulare, Ulothrix spp., and Spirogyra spp. Generally, the green macroalgae in the zinc-enriched streams included such species as Hormidium rivulare (98), Stigeoclonium tenue (114), Ulothrix moniliformis and Mougeotia sp. (109). Two hundred isolates comprising 87 species of Chlorophyta were obtained from the sites along the rivers that drain the ancient Cu and Pb mining region in Cornwall (U.K.) (100,101). They were classified mainly into the following taxonomic groups: Ulotrichales (Microspora stagnorum, M. pachyderma, M. willeana, M. tumidula, Ulothrix spp., Hormidium spp., Stigeoclonium tenue, Cylindrocapsa, Stichococcus spp.); Zygnematales (Mougeotia, Spirogyra), Volvocales and Tetrasporales (Chlamydomonas, Gleococcus, Chlamydocapsa, Asterococcus), Chlorococcaceae (Chlorella, Hypnomonas, Pleurococcus, Trochiscia); Oocystaceae (Oocystis, Pseudococcomyxa), and Scenedesmaceae (Ankistrodesmus, Selenastrum, Scenedesmus). As reported by Whitton (1), abundant growths of the filamentous green alga Stigeoclonium combined with the complete absence of Cladophora was often associated with high concentrations of heavy metals combined with relatively high nutrient concentrations. Conversely, massive growths of Cladophora glomerata usually indicated that heavy metals were present in flowing waters only at very low concentrations (14). The algal survey in three Japanese rivers polluted with Cu and Zn showed (103) that green algae were represented by different species of Chlorella, Oocystis, Scenedesmus, Chlamydomonas, Ulothrix, Klebsormidium, Stigeoclonium, Stichococcus, and Cosmarium. In the Cu- and Zn-polluted river Miyata in Japan, among a very limited number of periphytic algal flora Stigeoclonium aestivale and Oocystis lacustris (Chlorophyta) also appeared (108). The green microalgae Chlorella fusca var. vacuolata and Scenedesmus acutiformis were isolated from lake water in Canada, contaminated with Cu and Ni (96). The natural algal assemblages from the lakes (USA) contaminated with arsenic consisted of planktonic species of Chlorophyta: Chlorella sp., as well as mucilage sheath forming Chlamydocapsa bacillus and Chlamydocapsa cf. petrify (115). The isolated algae showed different sensitivity to various chemical species of arsenic, which can play a role in the selection of the phytoplankton communities. Recently, the periphytic green alga Oocystis nephrocytioides was identified as dominant in the freshwater communities exposed to high Cu concentrations (105). An abundant growth of Stigeoclonium sp. (104) accompanied by minor growths of a few species of green microalgae Scenedesmus spp., Chlorella sp., and Dictyococcus sp. (Pawlik-Skowron´ska, personal communication) was observed in the fast-flowing water of a mine drainage stream of elevated Zn, Cd, and Pb concentrations in southern Poland.

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4.3

Diatoms

In addition to Cyanophyta and Chlorophyta, some Bacillariophyceae (Chromophyta) also were reported in heavy metal–polluted freshwater. In slower flowing, metal-rich waters, planktonic microalgae included several diatom genera, e.g., Eunotia, Pinnularia, Navicula, and Synedra (109). In the mountain stream in Canada (water pH about 7), subjected to Zn, Cu, and Cd contamination, Deniseger et al. (102) reported the occurrence of such periphytic species of Bacillariophyceae as Achnanthes minutissima, Achnanthes microcephala, Synedra filiphormis, and Nitschia sp. The same diatom species were present both above and below heavy metal mining operations; however, at the downstream contaminated site the virtual absence of Chlorophyta was observed as compared with upstream control site. The Bacillariophyceae isolated from three Japanese rivers polluted with Cu and Zn were represented by 20 species belonging to the following genera: Achnanthes, Nitschia, Synedra, Gomphonema, Navicula, Pinnularia, Eunotia, and Cymbella. In the water of the urban river Miyata running through the abandoned copper mining region in Japan (108), among the very restricted periphyton species three diatoms were present: Achnanthes minutissima, Nitschia palea, and Surirella angustata. The river water was alkaline (pH was approximately 8) and polluted with Cu (0.07–0.22 mg/L) and Zn (0.13–0.55 mg/L). Achnanthes minutissima was also reported to be dominant in other copper-polluted rivers in Japan (113). As can be seen, all of the above-mentioned diatoms have been classified as the Naviculales order, and independently of the geographic location the same diatom species appeared in the heavy metal–polluted waters. However, it has been stated (116) that the differentiation in species composition can be associated with the local environmental conditions, such as pH of the water. For example, among diatoms, domination of Eunotia spp. in the acid streams and rivers is striking (117,118), and for the higher pH, Achnanthes spp., Nitschia spp., Navicula spp., and Gomphonema spp. are typical (117). Species richness and total biomass in freshwaters have been shown to diminish with the pH decrease (116,119). As reported by Kwandrans (120), in the acidic stream (U.K.) polluted with coal mine wastes (pH 2.6–3.3) the number of species was very limited and the algal community consisted only of Klebsormidium rivulare, Microthamnion strictissimum (Ulotrichales), Eunotia exigua, Pinnularia acoricola, Nitschia sp. (Bacillariophyceae), mucilage sheath–forming Gleochrysis turfosa (Chrysophyceae), and the green microalgae Euglena mutabilis and Chlamydomonas acidophila. However, among the diatoms inhabiting freshwaters of various pH values only a few taxa proved resistant to heavy metals; i.e., Achnanthes microcephala, Achnanthes minutissima, Achnanthes linearis, Eunotia exigua, Synedra filiformis, and Nitschia sp. (102,105,108,113,121–123).

Freshwater Algae

4.4

71

Metal Effects on Algal Communities

As reported by many authors (100,102,105,123,124), both algal abundance and algal diversity as measured by the number of species were reduced in high metal-polluted sites. Associations of species that were evident in the field samples were correlated with water metal levels (Table 3). For Cu, Pb, and Fe, the metal concentrations in algae were positively correlated with water metal levels (100). As observed by Foster (100), the degree of floristic stability correlated with the degree of heavy metal pollution rather than the polluting metal per se. This was shown by the close similarity between the algal flora on the copper-polluted river and the lead-polluted river. At high metal pollution (Cu, Cd, Pb, and Zn) an association of filamentous algae of the genus Microspora (Microspora stagnorum, M. pachyderma, M. willeana) accompanied by the microflora of Pseudococcomyxa adhaerens, Microthamnion kutzingianum, and Chlamydomonas vulgaris was observed (100). In the sites of low metal pollution the typical filamentous flora was Spirogyra, Mougeotia, Microspora floccosa, and the variety of planktonic species belonging to Scenedesmaceae, Oocystaceae, and Desmidiales. The moderately polluted sites had flora consisting of Zygogonium ericetorum, Microspora tumidula, Hormidium spp., Microthamnion spp., Stigeoclonium tenue, and many Volvocales and Tetrasporales species. As follows from many papers (97,98,100,102), there are no species invariably indicative of heavy metal pollution; however, it seems that there are taxonomic groups of green algae, e.g., Ulotrichales (Microspora, Hormidium, Stigeoclonium), mucilaginous Chlorophyceae (Chlamydomonas, Chlamydocapsa, Oocystis), and some pennate diatoms (Achnanthes, Nitschia, Navicula) abundantly occurring in the freshwater sites encompassing large ranges of heavy metal concentrations. These algae may be characterized by their abilities to adapt to high metal concentrations but cannot be estimated as strictly indicative of them. For instance, various Stigeoclonium species were found in the toxic metal–free areas (125,126) as well as in the streams and rivers polluted with various heavy metals (104,108,127,137). It is still difficult to make generalizations about the influence of heavy metal pollution on algal species and community composition. The individual situations in nature are very diverse, ranging from lakes to fastflowing streams, from one metal to a mixture of many, from nutrient-poor to nutrient-rich water of different pH values and load of suspended matter. In heavy metal–polluted, fast-flowing waters, periphyton communities well attached to the substrate dominate and the sort of substrate can be also responsible for species composition of the periphyton community (97). The water pH is a very important factor influencing directly (on cell physiology) or indirectly (on water chemistry) the algal development (116). It is also known that the algal community structure

72

TABLE 3 Algal Associations from the 64-Species Analysisa Metal levels

Low

Moderate

Moderate-high

High or variable

Chlorella fusca var. vacuolata ⬍ 4 µm Ulothrix variabilis Batrachosperum vagum Staurastrum punctulatum Ankistrodesmus falcatus var. mirabilis Pleurococcus vulgaris Spirogyra laxa? Chlorella fusca var. fusca

Hypnomonas lobata Hormidium rivulare Chlorella vulgaris var. vulgaris Chlamydomonas hebes Mougeotia gracillima? Zygogonium ericetorum Chlamydomonas debaryana Chlamydomonas botryopara

Microthamnion ku¨tzingianum Chlorella saccharophila var. ellipsoidea Pleurococcus rufescens Stigeoclonium tenue Oedogonium 7 µm

Pseudococcomyxa adhaerens Microspora pachyderma Microspora stagnorum Mougeotia parvula?

Microthamnion strictissimum Hormidium scopulinum Hormidium pseudostichococcus Hypnomonas chlorococcoides var. incrassata

Pawlik-Skowron´ska and Skowron´ski

Spirogyra nitida Spondylosium pygmaeum Chlamydomonas debaryana var. micropapilli Closterium striolatum

Low-moderate

Low-moderate Cu, high Zn, Pb, Fe

Microspora tumidula Chlamydomonas globosa Euglena gracilis

Approximate metal concentrations (mg L⫺1) Low Cu Pb Zn Fe

⬍0.1 ⬍0.1 ⬍0.5 ⬍0.2

Moderate

High

0.1–0.2 0.1–0.2 0.5–1.0 0.2–2.0

0.2–0.4 0.2–0.4 1.0–2.5 ⬎2.0

Freshwater Algae

Ankistrodesmus falcatus var. tumidus Scenedesmus serratus Chlamydomonas terricola Hypnomonas chlorococcoides Chlamydomonas heterogama Asterococcus limneticus Scenedesmus armatus Oocystis parva a

Species with similar loadings for two or more principal factors have been grouped according to their distributions at the sites. The approximate metal concentrations represented by the descriptive assignment are also given (mg L⫺1 total concentrations). Source: Ref. 100.

73

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of freshwaters depends on the season of the year (126), and spring or summer communities may react differently to heavy metal contamination. Dissimilarity index reported by Deniseger et al. (102) for the heavy metal–polluted stream community was high in summer and low in spring. However, there is evidence (128) that generally the dissimilarity of algal periphyton communities in the stream mesocosm increased with the increase of Zn concentration. Despite the complex condition in the freshwater ecosystems, the heavy metal impact is usually reflected by qualitative and quantitative changes in the algal community composition (129,130). Taking into account the above-mentioned data, it seems possible to distinguish some taxonomic groups that occur more often in heavy metal–polluted sites, i.e., some filamentous green algae belonging to Ulotrichales and Chaetophorales, microalgae like Chlorococcaceae, Oocystaceae, Scenedesmaceae, Volvocales, some Cyanophyta occurring in the alkaline or neutral waters, and some pennate diatoms belonging to Naviculales, e.g., Achnanthes spp. in the neutral/ alkaline waters and Eunotia exigua in the acidic environment. 4.5

Development of Metal-Resistant/Tolerant Communities

Although many biotic environmental factors are responsible for the algal community composition, heavy metals can act as agents of selection leading to establishment of the metal-tolerant ecotypes under characteristic local conditions. It was observed in nature, where metal-tolerant algae were isolated from the metalenriched sites (e.g., Zn-tolerant filamentous green algae H. rivulare and Stigeoclonium tenue (98,127), Zn-resistant coccoid green algae Oocystis elliptica and the blue–green alga Phormidium (128), Cd-tolerant Chlamydomonas spp. (101), Cu-tolerant Oocystis nephrocytioides (105,146), As-tolerant Chlorella sp., Chlamydocapsa spp. (115), and under the laboratory conditions Zn-tolerant Anacystis nidulans (131) and Selenastrum capricornutum (132). The tolerance levels (determined as heavy metal sensitivity of algal photosynthesis) of freshwater Bacillariophyceae, Charophyceae, and Chlorophyceae showed significant positive correlations with Cu concentrations in the field, whereas for Cd and Zn the correlations were not so clear. As shown in Fig. 2, the algae (excluding blue–green) isolated from the metal polluted sites were less sensitive to the polluting metals than those from the unpolluted sites (103). Pollutants such as heavy metals may affect communities because of the exclusion of sensitive species, or by the replacement with tolerant or opportunistic life forms. In aquatic environments subjected to pollution stress, increased tolerance has been found in the populations of algae. The concept of pollution-induced community tolerance (PICT) proposed by Blanck et al. (133) states that exposure of a community to a toxicant above its effect threshold exerts a selection pressure

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FIG. 2 Relative photosynthesis (%) of Achnanthes minutissima isolated from a polluted site (left) and an unpolluted site (right) of the River Miyata. (From Ref. 103.)

TABLE 4 Structural and Physiological Effects of Cu on Microalgal Communities in Freshwater Enclosures Time (days)

Culow

Cuhigh

⬎2

No effects on biomass, photosynthesis activity, or short-term tolerance. Small change in species composition.

2–14

Decrease in photosynthesis activity.

14–20

Increased tolerance for copper in short-term test. Cotolerance for zinc.

Very strong reduction in biomass and photosynthesis activity. Increased shortterm tolerance. Strong change in species composition. Continuation of the effects seen during the first 2 days. Increased biomass and photosynthesis activity. High tolerance for copper and cotolerance for zinc.

Source: Ref. 135.

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on sensitive species or individuals leading to their exclusions. As a result, the community will consist of more tolerant life forms and the overall tolerance of the community will increase as a consequence of this toxicant-induced succession. This has been already validated for the periphyton communities under arsenate stress (134), for the microalgal communities under copper stress (Table 4), and for the stream epilithon exposed to Zn (136). Also, in a field study of the heavily zinc-contaminated stretch of the river Dommel in Belgium (137), the microbenthic algal communities showed high tolerance to Zn. These elevated tolerance levels might be due to a long-term selection, favoring tolerant organisms and leading to a tolerant community as predicted by the PICT concept (134). As reported recently by Paulsson et al. (123), the species composition of riverine algal community changed slightly at low Zn concentrations but was strongly affected at high Zn stress (Fig. 3). In the case of copper effects on the structure

FIG. 3 Periphyton species richness after long-term exposure to zinc. Species richness represents the number of algal taxa or groups of taxa found in any of the 50 fields counted on a glass disc (n ⫽ 2–3) at each concentration. Zinc concentrations are based on analyzed total zinc in aquaria water. Bars represent standard deviations. The horizontal solid line indicates the mean baseline of no effect. The solid regression line is based on values between 0.28 and 25 µM of zinc (n ⫽ 17; P ⬍ 0.001). The dotted lines indicate 95% confidence limits for the regression. (From Ref. 123.)

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of freshwater algal periphyton communities, it has been also stated (105) that the Cu-induced structural impacts on the periphyton communities can be evidenced as an increased tolerance to copper. As shown in Table 5, the long-term exposure of the periphyton communities of the river Glatt to Cu caused a dominance shift from Cyanophyceae to Chlorophyta, while the relative abundance of Bacillariophyceae was stable independent of the Cu treatments. Each class of the studied algae showed broad interspecific differences in sensitivity to Cu. Among Cyanophyceae the most tolerant was Pseudoanabaena catenata (dominant at 1 µM Cu) while among Chlorophyta the most tolerant was Oocystis nephrocytioides, whose relative abundance increased from less than 1% (in control) to 56% in 5-µM Cu treatments (105). Of the diatoms again Achnanthes was the dominant genus at 5 µM Cu. As shown in Table 5, upon long-term Cu exposure the most sensitive algal species were eliminated and the tolerant-species dominated. The

TABLE 5 Relative Abundance (%) of Algal Taxa in Periphyton Communities After 12 Weeks of Exposure to Copper a Copper conc. (µM)

Chlorophyta Cosmarium sp. Geminella interrupta Mougeotia sp. Oedogonium sp. Oocystis nephrocytioides Phacotus lenticularis Scenedesmus spp. Others Bacillariophyceae Achnanthes sp. Fragilaria sp. Tabellaria fenestrata Others Cyanophyceae Gomphospheria sp. Microcystis sp. Oscillatoria sp. Pseudoanabaena catenata

Control

0.05

0.1

0.5

1

5

⬍1 7 3 ⬍1 ⬍1 1 4 9

⬍1 3 6 14 14 ⬍1 5 3

1 18 7 2 8 1 9 7

1 11 11 4 22 1 2 13

4 3 1 0 20 1 0 13

0 0 0 0 56 5 0 13

6 2 3 3

6 1 2 8

2 2 4 5

4 1 3 11

0 3 3 15

14 2 1 3

10 27 25 ⬍1

7 6 21 3

24 3 0 8

3 6 0 8

2 0 0 34

0 0 5 0

a Values represent means of three measurements done with composite samples taken from each of three treatment replicates. Source: Ref. 105.

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observed species selection and structural shifts of the community due to the presence of pollutant also support the concept of the pollution-induced community tolerance. However, the relationship between heavy metal concentration and community response, such as shifts in the tolerance to metals, is confounded by the influence of chemical speciation on metal bioavailability. Metal species formation is influenced by various local physicochemical conditions such as acidity, salinity, inorganic and organic ligands, and particulate matter (138).

5

VARIETY OF MECHANISMS OF METAL RESISTANCE/ TOLERANCE

The ability of algae to survive and reproduce in the heavy metal–polluted waters has not been explained so far and may rely on intrinsic properties of the organisms as well as the physical and chemical nature of the habitat. It can also depend on genetic adaptation over the extended time periods by mutation, genetic exchange, selection, etc., or on changes in the organism physiology resulting from metal exposure (139). Metal resistance can be metal-specific, e.g., the green alga Spirogyra sp. isolated from the Cu- and Zn-polluted river showed specific resistance only to Zn, while some pennate diatoms and the green alga Klebsormidium klebsii showed specific resistance only to Cd (103). Since metal pollution is rarely confined to one metal, combined resistance to several metals (multiple resistance, cotolerance) can be expected to occur among algae from such environments. Cotolerance to zinc and cadmium was evident in some freshwater algal populations, and this resistance can result from the presence of both metals in the polluted sites (98,101). The Cu-resistant algae studied by Foster (101) were significantly Pb-resistant; however, the Pb-resistant algae were Cu-sensitive. Cu-tolerant Bacillariophyceae and Cd-tolerant Chlorophyceae isolated from the polluted river tended also to be Zn-tolerant (103). Freshwater phytoplankton communities exposed to the increased copper concentrations also showed increased tolerance for zinc (135). Algal periphyton communities exposed to Cu (105) displayed an increased cotolerance to Zn, Ag, and Ni. This phenomenon has been explained (101) as resistance achieved by a mechanism to cope with excessive levels of another metal. However, there is no firm evidence that resistance to different metals can depend on the same genes. Within algal populations surviving under the chronic metal exposition, different evolutionary strategies, conveying different types and degrees of metal resistance, appear to coexist. Regardless of whether it is constitutive or induced upon exposure, metal resistance is usually a heritable phenomenon. According to Stokes (140), metal resistance is used to describe the algal species in which all populations are normally able to tolerate metals, while algal tolerance occurs in the species of widespread distribution, with populations in the unpolluted envi-

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ronments being less metal tolerant. It is assumed that tolerant populations are genetically distinct from those that are not. However, tolerance and resistance are arbitrary terms often used interchangeably (141). Following the Levitt definition (142), heavy metal resistance can be achieved in two ways: avoidance or tolerance. Avoidance is defined as the ability to prevent excessive metal uptake. The tolerance is an organism’s ability to cope with metals that are excessively accumulated within its body. 5.1

Metal Avoidance

There are some assumptions that extracellular organic material can be involved in the enhanced heavy metal resistance of some strains of algae. For instance, the extracellular material produced by Gleocystis gigas (⫽ Chlamydomonas ampla) (143) and Chlorella vulgaris (144) reduced cupric ion activity and thus metal toxicity. Similar observation has been made recently for the Cu-resistant green alga Oocystis pusilla (145). However, Cu complexation was not essential for Cu tolerance of another species Oocystis nephrocytioides (146). High levels of cell wall–associated copper have been reported for other Chlorophyta (147); in the freshwater green algae such as Cladophora, Chlorella, Scenedesmus, Spirogyra, and Chara, cell walls played important role in heavy metal binding (45,148,149). The surface metal binding may protect cells to some extent from the toxic effects by reducing the concentration of free metal cations in surrounding water. As mentioned before, the surface of algae contains a number of functional groups with high affinity for metal ions. Some of these groups are inert adsorption sites, not involved in metal transport across the cell membrane. Decreased internal accumulation of heavy metals has been also proposed as a mechanism of resistance in certain algae. In the green alga Scenedesmus acutiformis, Cu transport across plasmalemma was extremely slow compared with a sensitive strain (150). Also, the Cd resistance of Euglena gracilis (151) can rely on the modification of membrane permeability. The decreased affinity of membrane permeases to Cd ions reduced the metal uptake. A similar explanation has been proposed by Rai et al. (152) for the Cu tolerance of the blue–green alga Anabaena doliolum. The heavy metal resistance of the green alga Dunaliella acidophila resulted from the reduced uptake of metal cations due to the positive ζ potential of the plasma membrane (153). More recently, a low Ni uptake in acid-tolerant Chlorella vulgaris, brought about by a change in the membrane potential and permeability, was offered as a resistance mechanism to metals (194). It is in agreement with the observation (155) that the Cd, Zn, Cu, and Ni resistance in the mutant strains of Chlamydomonas reinhardtii was associated with reduced active metal uptake. It has also been shown that the development of Cu resistance in Euglena can be related to the changes in structural carbohydrates in the cell surface (156).

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Metal exclusion as another mechanism of Cu and Zn resistance both in green and blue–green algae has been proposed (157,158,159). In prokaryotic blue–green algae, however, the metal resistance mechanism would rely on the plasmid-encoded metal efflux (160,161), as in the Cd-resistant bacteria Pseudomonas putida (162). 5.2

Metal Tolerance

Metal tolerance relies on detoxification mechanisms operating within algal cells. The possibility of internal detoxification in freshwater algae has received more attention recently. 5.2.1

Production of Intracellular Metal-Binding Compounds

A common response of algae upon exposure to heavy metals is the synthesis of intracellular metal-binding compounds, which may function in detoxification of these metals (150,163–165). Eukaryotic algae synthesize sulfur-rich oligopeptide phytochelatins similar to those found in higher plants (166–168). Phytochelatins are produced in algal cells intracellularly in response to several heavy metals and metalloids, though Cd appears to be the most effective inducer. They are small peptides typically having the structure (γ Glu-Cys)n-Gly, n ⫽ 2–11 (163), synthesized enzymatically. Phytochelatins can complex metals intracellularly via sulfhydryl groups of cysteine. As shown in Table 6, a number of metals, such as Cd, Cu, Zn, Pb, Hg, and Ag, induce phytochelatins in different taxons of freshwater algae (27,63,104,166–174). However, there have been reported some interspecies differences (166,174) concerning both total levels of these peptides and particular oligomer concentrations (Fig. 4) as in marine phytoplankton (175,176). Most studies concerning the phytochelatin synthesis in freshwater algae exposed to heavy metals were carried out under laboratory conditions (27,166,168,170,171,174). However, recently phytochelatins were identified and determined in the field population of freshwater algae in the metal-contaminated lakes in Switzerland (173) and in the field populations of Stigeoclonium sp. growing abundantly in a mine drainage stream of elevated Zn, Pb, and Cd levels in Poland (104). Intracellular binding of Cd by phytochelatins in some algal cells was stated in vivo (166,167), and a considerable proportion of incorporated Cd in the unicellular green alga Chlamydomonas reinhardtii was sequestered by complex formation with phytochelatin oligomers in both cytosol and chloroplasts (172). Intracellular metals were found to be of magnitude order higher than phytochelatin levels in the algae inhabiting the heavy metal–polluted waters (104,173). It suggests that phytochelatins cannot be the only tolerance mechanism; however, they still may be a very important part of a detoxification process as the first line of defense together with glutathione (GSH). GSH plays a role as a precursor of phytochelatin synthesis; however, there are some data showing that GSH may

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TABLE 6 Freshwater Algae that Produce Phytochelatins in Response to Heavy Metals Organism

Metals

Ref.

Navicula pelliculosa (Bacillariophyceae) Cd 166 Fragillaria crotonensis (Bacillariophyceae) Cd 166 Thalassiosira pseudonana (Bacillariophyceae) Cd, Zn, Cu 63 Euglena gracilis (Euglenophyceae) Cd 166 Chlorella fusca (Chlorophyceae) Cd, Pb, Zn, Ag, Cu, Hg 166 Chlorella sp. (Chlorophyceae) Cd 170 Scenedesmus acutiformis (Chlorophyceae) Cd, Pb, Zn, Ag, Cu, Hg 166 Scenedesmus quadricauda (Chlorophyceae) Cd 168 Scenedesmus subspicatus (Chlorophyceae) Cu 173 Chlamydomonas reinhardtii (Chlorophyceae) Cd 166, 167, 172 Chlamydomonas reinhardtii (Chlorophyceae) Hg 167 Stichococcus bacillaris (Chlorophyceae) Cd 166, 171 Stichococcus bacillaris (Chlorophyceae) Pb 27 Monoraphidium minutum (Chlorophyceae) Cd 166 Stigeoclonium sp. (Chlorophyceae) Metal mixture 104 (Zn, Pb, Cd) Stigeoclonium tenue (Chlorophyceae) Metal mixture 104 (Zn, Pb, Cd) Bumilleriopsis filiformis (Xanthophyceae) Cd 166 Vaucheria compacta (Xanthophyceae) Cd 174 Vaucheria debaryana (Xanthophyceae) Cd 174 Mixed lake phytoplankton Metal mixture 173 (Cu, Cd, Zn)

sometimes play an essential role in heavy metal detoxification. In the case of metals such as Hg, Ag, and Pb, which create mercaptide complexes of high stability (177,178), an excessive production of GSH besides phytochelatins was observed in some freshwater Chlorophyta (27,167). It is also likely that phytochelatins and probably GSH can serve in algae as a shuttle of metals to the vacuole, as has been demonstrated in the yeast and plant tissue (179,180), or to the cytoplasmic membrane where they may be exported from the cells with metal ions. The latter possibility was suggested for Cd and marine diatom (181) as well as for Pb and freshwater green alga Stichococcus bacillaris (27). It has recently been evidenced in higher plants and yeasts (182) that genes encoding the enzyme phytochelatin synthase mediate Cd detoxification in eukaryotes. In the case of prokaryotic blue-green algae like Synechococcus Cd and Zn tolerance involves intracellular binding of metals to ligands similar to some extent to eukaryotic metallothioneins (183–185). Class II cyanobacterial metallothionein, however, does not appear to play a role in Cu detoxification.

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FIG. 4 Effect of external Cd concentration on phytochelatin (PC) production in V. compacta (A) and V. debaryana (B) after 3-day exposure. Temperature of Cd exposure: (A) 16°C; (B), 18°C. In control cultures (Cd-free) PCs were not detected. (From Ref. 174.)

Freshwater Algae

5.2.2

83

Metal Compartmentalization

It has been reported for several algae (both prokaryotic and eukaryotic) that intracellular metal sequestration may involve precipitation within the specific intraprotoplast sites. The most frequently reported sites are polyphosphate granules in algae, which store inorganic phosphate and metabolically essential metals (92,148,186–188). The ability of polyphosphate bodies to accumulate a number of heavy metals like Fe, Zn, Cd, Cu, and Pb may serve as protection of algal cells from metal toxicity. The other cell compartment that plays a role in heavy metal sequestration in plant organisms is the vacuole (189). There are scarce data, however, concerning the accumulation of metals in algal vacuoles. Silverberg (190) reported Pb in the peripheral vacuoles of the filamentous green alga Stigeoclonium tenue in Pb exposed cells. X-ray microanalysis performed on cells of flagellate green alga Dunaliella bioculata revealed that Cd was associated with sulfur in vacuoles (191). Recently, the rapid transfer of zinc to vacuole of the giant freshwater alga Chara corralina was also reported (45). 5.2.3

Metal Transformations

Microorganisms can carry out chemical transformations of heavy metals such as oxidation, reduction, alkylation, and dealkylation, and some of these have been proposed as mechanisms of metal resistance (192,193). However, there are only a few data showing the involvement of algae in metal transformation. For instance, in the Hg-resistant Chlorella, enzymatic reduction and volatilization of Hg were reported (158). Debutylation of toxic tributyltin (53) and demethylation of trimethyllead were reported in the freshwater chlorophyte Ankistrodesmus falcatus. Biomethylation and excretion of arsenic have been reported in the Asresistant Chlorella vulgaris (194). These scarce data do not give any idea about the role of metal biotransformations in the development of heavy metal resistance in algae. There are reports showing that organic-metal complexes of Sn, Pb, or As can be more toxic for algae than inorganic forms of heavy metals (50,115,195). The phenomenon of heavy metal tolerance in freshwater algae is still not well recognized and can rely on several mechanisms operating simultaneously in the exposed organism. REFERENCES 1. BA Whitton. Algae as monitors of heavy metals in freshwaters. In: LE Shubert, ed. Algae as Ecological Indicators. London: Academic Press, 1984, pp. 257–280. 2. PGC Campbell, PM Stokes. Acidification and toxicity of metals to aquatic biota. Can J Fisher Aquat Sci 42:2034–2049, 1985. 3. JT Trevors, GW Stratton, GM Gadd. Cadmium transport, resistance and toxicity in bacteria, algae and fungi. Can J Microbiol 32:447–464, 1986. 4. J Vymazal. Occurrence and chemistry of zinc in freshwaters—its toxicity and bio-

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148. M Pedersen, GM Roomans, M Andre¯n, A Lignell, G Lindahl, K Wallstro¨m, A Forsberg. X-ray microanalysis of metals in algae—a contribution to the study of environmental pollution. Scan Electr Microsc 2:499–509, 1981. 149. S Saygideg˘er. Bioaccumulation and toxicity of zinc in Spirogyra fluviatilis hilse (Chlorophyta). Wat Air Soil Pollut 101:323–331, 1998. 150. PM Stokes. Uptake and accumulation of copper and nickel by metal-tolerant strains of Scenedesmus. Verh Internat Verein Limnol 19:2128–2137, 1975. 151. A Bariaud, JC Mestre. Heavy metal tolerance in a cadmium-resistant population of Euglena gracilis. Bull Environ Contam Toxicol 32:597–601, 1984. 152. LC Rai, N Mallick, JB Singh, HD Kumar. Physiological and biochemical characteristics of a copper tolerant and a wild type strain of Anabaena doliolum under copper stress. J Plant Physiol 138:68–74, 1991. 153. H Gimmler, B Treffny, M Kowalski, U Zimmermann. The resistance of Dunaliella acidophila against heavy metals: the importance of the zeta potential. J Plant Physiol 138:708–716, 1991. 154. PK Rai, N Mallick, LC Rai. Effect of nickel on certain physiological and biochemical behaviors of an acid tolerant Chlorella vulgaris. BioMetals 7:193–200, 1994. 155. JM Collard, RF Matagne. Cd2⫹ resistance in wild-type and mutant strains of Chlamydomonas reinhardtii. Environ Exp Bot 34:235–244, 1994. 156. J Bonaly, E Brochiero. Cell-surface changes in cadmium-resistant Euglena: studies using lectin-binding techniques and flow cytometry. Bull Environ Contam Toxicol 52:54–60, 1994. 157. PL Foster. Copper exclusion as a mechanisms of heavy metal tolerance in a green alga. Nature 269:322–323, 1977. 158. LF De Filippis, CK Pallaghy. The effect of sub-lethal concentrations of mercury and zinc on Chlorella. III. Development and possible mechanisms of resistance to metals. Z Pflanzenphysiol 79:323–335, 1976. 159. SK Verma, HN Singh. Evidence for energy-dependent copper efflux as a mechanism of Cu2⫹ resistance in the cyanobacterium Nostoc calcicola. FEMS Microb Lett 84:291–294, 1991. 160. SP Singh, K Pandey. Cadmium-mediated resistance to metals and antibiotics in a cyanobacterium. Mol Gen Genet 187:240–243, 1982. 161. N Takamura, F Kasai, MM Watanabe. Unique response of Cyanophyceae to copper. J Appl Phycol 2:293–296, 1990. 162. K Kawai, H Horitsu, K Hamada, M Watanabe. Induction of cadmium-resistance of Pseudomonas putida GAM-1. Agric Biol Chem 54:1553–1555, 1990. 163. E Grill, EL Winnacker, MH Zenk. Phytochelatins: the principal heavy-metal complexing peptides of higher plants. Science 230:674–676, 1985. 164. T Nagano, M Miwa, Y Suketa, S Okada. Isolation, physicochemical properties, and amino acid composition of a cadmium-binding protein from cadmium-treated Chlorella ellipsoidea. J Inorg Biochem 21:61–71, 1984. 165. DJ Gingrich, DN Weber, CF Shaw, JS Garvey, DHA Petering. Characterization of a highly negative and labile binding protein induced in Euglena gracilis by cadmium. Environ Health Perspect 65:77–85, 1986. 166. W Gekeler, E Grill, EL Winnacker, MH Zenk. Algae sequester heavy metals via synthesis of phytochelatin complexes. Arch Microbiol 150:197–202, 1988.

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167. G Howe, S Merchant. Heavy metal-activated synthesis of peptides in Chlamydomonas reinhardtii. Plant Physiol 98:127–136, 1992. 168. GN Reddy, MN Prasad. Characterization of cadmium binding protein from Scenedesmus quadricauda and Cd toxicity reversal by phytochelatin consulting amino acids and citrate. J Plant Physiol 140:156–162, 1992. 169. T Skowron´ski, JA De Knecht, AP van Beem, RA Broekman, J Simons, JAC Verkleji. Cadmium accumulation and detoxification in Vaucheria compacta (Xanthophyceae). Proceedings of International Conference on Heavy Metals in the Environment, Toronto, 1993, CEP Consultants, Ltd., Edinburgh, pp. 312–315. 170. D Kaplan, YM Heimer, A Abeliovich, PB Goldsbrough. Cadmium toxicity and resistance in Chlorella sp. Plant Sci 109:129–137, 1995. 171. T Skowron´ski, J Pirszel, B Pawlik-Skowron´ska. Phytochelatin synthesis—response to cadmium in the green microalga Stichococcus bacillaris. Proceedings of International Conference on Heavy Metals in the Environment, Hamburg, CEP Consultants, Ltd., Edinburgh, pp. 224–227, 1995. 172. K Nagel, U Adelmeier, J Voigt. Subcellular distribution of cadmium in the unicellular green alga Chlamydomonas reinhardtii. J Plant Physiol 149:86–90, 1996. 173. K Knauer, BA Ahner, HB Xue, L Sigg. Metal and phytochelatin content in phytoplankton from freshwater lakes with different metal concentrations. Environ Toxicol Chem 17:2444–2452, 1998. 174. T Skowron´ski, JA De Knecht, J Simons, JAC Verkleji. Phytochelatin synthesis in response to cadmium uptake in Vaucheria (Xanthophyceae). Eur J Phycol 33:87– 91, 1998. 175. GH Wikfors, A Neeman, PJ Jackson. Cadmium-binding polypeptides in microalgal strains with laboratory-induced cadmium tolerance. Mar Ecol Prog Ser 79:163– 170, 1991. 176. BA Ahner, S Kong, FMM Morel. Phytochelatin production in marine algae: I. An interspecies comparison. Limnol Oceanogr 40:649–657, 1995. 177. W Stricks, IM Kolthoff. Reactions between mercuric mercury and cysteine and glutathione. Apparent dissociation constants, heats and entropies of formation of various forms of mercuric mercapto-cysteine and -glutathione. J Am Chem Soc 75:5673–5681, 1953. 178. RK Mehra, VR Kadati, R Abdullah. Chain length-dependent Pb(II)-coordination in phytochelatins. Bioch Biophys Res Commun 215:730–736, 1995. 179. DF Ortiz, L Kreppel, DM Speiser, G Scheel, G McDonald, DW Ow. Heavy metal tolerance in the fission yeast requires an ATP-binding cassette-type vacuolar membrane transporter. EMBO J 11:3491–3499, 1992. 180. DE Salt, WE Rauser. MgATP-dependent transport of phytochelatin across the tonoplast of oat roots. Plant Physiol 107:1293–1301, 1995. 181. JG Lee, BA Ahner, FMM Morel. Export of cadmium and phytochelatin by the marine diatom Thalassiosira weissflogii. Environ Sci Tech 30:1814–1821, 1996. 182. S Clemens, EJ Kim, D Neumann, JI Schroeder. Tolerance to toxic metals by a gene family of phytochelatin synthases from plants and yeasts. EMBO J 18:3325– 3333, 1999. 183. RW Olafson. Physiological and chemical characterization of cyanobacterial metallothioneins. Environ Health Persp 65:71–75, 1986.

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184. NJ Robinson. Algal metallothioneins: secondary metabolites and proteins. J Appl Phycol 1:5–18, 1989. 185. A Gupta, BA Whitton, AP Morby, JW Huckle, NJ Robinson. Amplification and rearrangement of a prokaryotic metallothionein locus smt in Synechococcus PCG 6301 selected for tolerance to cadmium. Proc R Soc Lond 248:273–281, 1992. 186. TE Jensen, JW Rachlin, W Jani, B Warkentine. An X-ray energy dispersive study of cellular compartmentalization of lead and zinc in Chlorella saccharophila (Chlorophyta), Navicula incerta and Nitzchia closterium (Bacillariophyta). Environ Exp Bot 22:319–328, 1982. 187. BC Rana, HD Kumar. The toxicity of zinc to Chlorella vulgaris and Plectonema boryanum and its protection by phosphate. Phykos 13:60–66, 1974. 188. MF Fiore, JT Trevors. Cell composition and metal tolerance in cyanobacteria. BioMetals 7:83–103, 1994. 189. J Wang, BP Evangelou, MT Nielsen, GJ Wagner. Computer simulated evaluation of possible mechanisms for sequestering metal ion activity in plant vacuoles. Plant Physiol 99:621–626, 1992. 190. BA Silverberg. Ultrastructural localization of lead in Stigeoclonium tenue (Chlorophyceae, Ulotrichales) as demonstrated by cytochemical and X-ray microanalysis. Phycol 14:265–274, 1975. 191. H Heuillet, A Moreau, S Halpern, N Jeanne, S Puiseux-Dao. Cadmium-binding to a thiol-molecule in vacuoles of Dunaliella bioculata contaminated with CdCl2: electron probe microanalysis. Biol Cell 58:79–86, 1986. 192. JM Wood, HK Wang. Microbial resistance to heavy metals. Environ Sci Technol 17:583–590, 1983. 193. JJ Cooney. Interactions between microorganisms and tin compounds. In: PJ Craig, F Glockling, eds. The biological alkylation of heavy elements. London: Royal Society of Chemistry, 1988, pp. 92–104. 194. S Maeda, K Kusadome, H Arima, A Ohki, K Naka. Biomethylation of arsenic and its excretion by the alga Chlorella vulgaris. Appl Organom Chem 6:407–413, 1992. 195. JJ Cooney, S Wuertz. Toxic effects of tin compounds on microorganisms. J Ind Microbiol 4:375–402, 1989.

4 Salt Marshes Isabel Cac¸ador University of Lisbon, Lisbon, Portugal

Carlos Vale Research Institute for Fisheries and Sea Research, Lisbon, Portugal

1

SALT MARSHES

Salt marshes are present in moderate- to low-energy intertidal zones of macroand mesotidal estuaries and bays. These areas have a great ecological value for the ecosystem, namely in terms of nutrient regeneration, primary production, habitat for fish and birds as well as other wildlife, and as shoreline stabilizers. The tidal flooding causes a transport of dissolved and particulate material from the estuarine waters to the salt marshes, and consequently seawater constituents, like sodium chloride, are progressively incorporated in the sediments (1–3). These conditions favor the colonization of salt marshes by vascular plants (Fig. 1) belonging to a few cosmopolitan genera (4). The dominant factor determining the composition of salt marsh flora (5) is the ability to withstand a sediment environment characterized by high salinity. Halophytes that colonize these areas possess a well-developed aerenchyma system through which atmospheric oxygen is transported from the leaves to the roots. The oxygen that is not consumed in the roots 95

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FIG. 1 Pancas salt marsh: a pure stand of Spartina maritima.

during respiration diffuses into the surrounding sediment and forms an oxidized rhizosphere (6–8). These oxidative microenvironments are important to the plant because they may increase nutrient availability and act as a protective barrier to heavy metal uptake (9). Cities and industrialized areas were often installed in the proximity of salt marshes in estuaries and coastal lagoons. In these cases, flooding transports large quantities of contaminants in both dissolved and suspended particulate forms to the salt marsh areas. Anthropogenic metals are incorporated in the sediments, decreasing their availability in the water column. Several studies have proposed that salt marshes act as natural sinks for metals (10–12). However, a number of factors may cause postdepositional mobilization of metals from industrial and urban origin. Diagenetic reactions related to oxidation of organic matter that usually occurs in high levels in salt marshes and oxidation of metal sulfides in the rooting zone (13) lead to the mobilization of metals. Complex interactions between salt marsh plants and sediments result in the redistribution of metals in the sediment-root system (14). Metals are taken up by the roots and translocated to the above-ground parts of the plants; when the plants die the biosynthesized organic matter is oxidized and the metals returned to the sediments. Since roots accumulate larger quantities of metals (15,16), the interactions between belowground biomass and sediments are extremely effective and may have a strong influence on the form and concentrations of metals in the rhizosphere (7). Removal or substitution of degradation products by the flooding water that perco-

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lates the upper sediment layers during the flood tide (17) forces the export of material to the water column.

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THE TAGUS ESTUARY SALT MARSHES: A CASE STUDY

The Tagus estuary is one of the largest estuaries on the Atlantic coast of Europe, covering an area of 300 km2 at low tide and 340 km2 at extreme high tide (Fig. 2). The southern and eastern parts of the estuary contain extensive intertidal mudflat areas with the presence of Spartina maritima (Poales: Poaceae), Halimione portulacoides (Caryophyllales: Chenopodiaceae), and Arthrocnemum fruticosum (Caryophyllales: Chenopodiaceae). Contrary to many cases in Europe where pollutants from industrial regions are discharged into rivers and brought to the estuaries via the rivers, in the Tagus most pollutants are discharged directly into the estuary. The estuary receives effluents from about 2.5 million inhabitants living in the greater Lisbon area, together with the discharges from industries (chemicals, steelmaking, and shipbuilding). Previous studies showed that Tagus salt marshes incorporate large quantities of anthropogenic metals into the sediments. Concen-

FIG. 2 The Tagus estuary and the salt marsh area: sampling sites.

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trations of copper (Cu), zinc (Zn), cadmiun (Cd), and lead (Pb) in the upper sediment increase 3–12 times relative to preindustrial levels, which indicates that a substantial quantity of anthropogenic metals is incorporated into the sediment (18).

3

METAL CONTAMINATION OF THE TAGUS SALT MARSHES SEDIMENTS

Vascular plants in salt marshes are determinant to the dynamics of the estuarine ecosystem. Plants act as sediment traps, facilitating an important settling of suspended estuarine material and their associated metals, and influence the retention and the accumulation processes of metals in salt marsh sediments (18). Several studies point to metal contamination of surface sediments from the salt marsh areas in the Tagus estuary (12,18). Higher contaminations were found close to the anthropogenic sources, highlighting the role of salt marshes in the retention of metals in this system (19). As in many other coastal systems located in the proximity of urban and industrial zones, metal contamination decreases with the sediment depth. The metal concentrations in surface sediments are often related to contamination of the coastal environment. An enrichment factor (EF) may be calculated by the ratio between metal concentration in the first sediment layer (0–5 cm) and in deeper sediment layers (45–55 cm). This depth is below the plant roots, avoiding the influence of the roots on metal redistribution, and sediments were deposited before the industrial period in Portugal (18). Consequently, the EF represents the increment of metal concentrations in the sediments due to environmental contamination (20). The calculated EFs for Zn, Pb, Cu, and Cd in the Tagus salt marshes are present in Fig. 3. The higher values were registered in Corroios (C), Talaminho (T), and Rosa´rio (R), which are the salt marshes nearby metal anthropogenic sources. Lower values were found in areas P and B located in the Natural Reserve of the Tagus estuary. The calculation of EF allows assessing the retention of anthropogenic metals in the salt marshes along the estuary. In spite of the clear picture obtained, it should be stressed this calculation is a simplistic approach when changes of metal concentrations due to early diagenetic processes (19) are not taken into account. The sedimentary organic matter is considered to have an important role in the retention of metals in sediments. The relationships between metal concentrations and organic matter content, here estimated from the loss on ignition (LOI) at 550°C (20) are shown in Fig. 4. These figures were plotted with values from three sediment layers: surface (0–5 cm), subsurface (5–15 cm), and deeper sediment layers (45–55 cm). In general, the deeper sediment layer has lower organic matter content and lower metal concentrations than the upper two layers. Metal

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FIG. 3 Enrichment factors: heavy metal concentration (0–5 cm)/(45–55) cm (µg µg⫺1), sediment layer for Zn, Pb, Cu, and Cd in Pancas (P), Barco (B), Sarilhos Pequenos (S), Rosa´rio (R), Paio Pires (p), Talaminho (T), and Corroios (C) salt marshes.

concentrations are not linearly correlated to LOI, which indicates that organic matter content is not the dominant factor for metal concentration in upper sediment layers of salt marshes. Though vegetated litter is the main source of organic matter to salt marshes (2), as well as an important vehicle of metals that have been uptaken by plants during their growing process to the sediment (11), anthropogenic inputs appear to be superimposed. Furthermore, metal concentrations and LOI in sediment sites colonized by Spartina maritima, Halimione portulacoides, and Arthrocnemum fruticosum did not show consistent differences, rein-

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FIG. 4 Relationships between the concentrations of Zn, Pb, Cu (µmol g⫺1), and LOI (%) in sediments from three layers: (0–5) (⫹), sub-superficial 5–15 ⋅ ), and 45–55 cm (䉭) of the Tagus salt marshes. cm (䊊

forcing the hypothesis of external sources in the geographic distribution of metal concentration in the upper sediment layers. The amount of metal extracted by the DTPA is assumed to be an estimation of the metal quantity potentially available to the plants (21). These amounts are relatively small in comparison to the total metal concentration, being lower than 10% for Cd and Cu and not exceeding 30% for Zn (7). Moreover, the extracted metal concentrations changed with the salt marsh. When metal fractions extracted by the DTPA solution are plotted against LOI (20), inverse relationships are obtained (Fig. 5). The best fitting curves have similar shapes and the general equation is [Me]extracted ⫽ A[LOI]b. This indicates that metals are more available in sediments impoverished in organic matter than in organically rich sediments. For example, available Zn was 10 times higher in sediments with LOI lower than 10% than in sediments containing more than 20%. These sediments showed low proportion of Pb, Cu, and Cd, while in sediments with low organic matter content the proportions were generally higher and varied with the metal. Otherwise, rooting sediments containing higher organic content have lower metal extracted by the DTPA solution.

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FIG. 5 Relationships of extractable Zn, Pb, Cu, and Cd (% to the total mass) and LOI (%) in sediments between roots (5–15 cm) of the Tagus salt marshes.

4

EFFECT OF PLANT ROOTS ON SEDIMENT CHEMISTRY

Roots may affect substantially the chemistry of sediment layers containing high below-ground biomass. The interaction between roots and sediments is exceedingly complicated, covering a wide range of chemical, physical, and biological processes (22). Plants take up nutrients and other minor and trace elements from the sediments, causing a flux of dissolved chemical species from the sediments toward the root surfaces. Furthermore, oxidants, exudates, and gases liberated by the roots change the surrounding sediment environment. Rooting sediments in the Tagus salt marshes are richer in organic matter and have a lower pH and higher redox potential than unvegetated bulk sediment (7) from nearby areas (Fig. 6). Vertical profiles of Zn, Pb, and Cu concentrations in vegetated areas differ in shape and concentrations of those found in nonvegetated sediments (Fig. 7). The major characteristic is the appearance of a subsurface maximum of several metals. Concentration peaks are formed at depths of higher rooting density, and levels in the upper 30 cm were higher than those in the corresponding layers of nonvegetated sediments. These differences are clearly related to the presence and activity of roots. Metal incorporation in salt marsh sediments is therefore substantially different from that observed in nonvegetated intertidal and subtidal areas (18). Apparently, the larger amounts of metals retained in the rooting sedi-

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FIG. 6 Redox potential (Eh) vs. pH, and Organic matter content (LOI) vs. pH in sediment between roots, corresponding mainly to the 5- to 15-cm layer (⫹), and in nonvegetated sediment (including sediment below roots) (*) at Corroios and Rosa´rio Tagus salt marsh.

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FIG. 7 Vertical profiles (mean SD, n ⫽ 3) of total Zn, Pb, and Cu concentrations (µg µg⫺1 dry weight) in sediments with (䉬) and without (ⱓ) vegetation from less contaminated site 1 and more contaminated site 2.

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FIG. 8 Distribution of Zn, Pb, and Cu among operationally defined fractions in three layers in vegetated and nonvegetated sediment from sites 1 and 2. Sediment fractions are exchangeable (■), carbonates ( ), oxides ( ), organic matter ( ), and residual (䊐).

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ment layers are not easily available to the plants. When compared with the sediments not colonized by vascular plants (12), sediments around roots have lower Pb, Zn, and Cu concentrations in mobile chemical forms (Fig. 8). In fact, the amount of metals taken up by the plants depends on their availability, which is a function of several factors—namely, organic matter, pH, redox potential, metal speciation, and root-sediment interactions (22)—which change drastically in the vicinity of the roots (rhizosphere). These results indicate higher percentages of metals unavailable to plants (12). 5

FORMATION OF METAL-RICH CONCRETIONS AROUND THE ROOTS

By releasing gaseous oxygen to the rhizosphere, vascular plants modify dramatically the sediment characteristics in the areas colonized by salt marsh vegetation.

FIG. 9 Concentrations of Zn, Pb, Cu, and Cd in rhizoconcretions (0) obtained with sieves of 2, 1, 0.5, and 0.25 mesh size and in sediment between roots (-).

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(A)

(B) FIG. 10 (A) Rhizoconcretions on the roots of Aster tripolium in the Tagus estuary. (B) Enlarged view of a concretion showing that root remains in its interior.

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The presence of oxygen at the root-sediment interface creates local oxidizing conditions in otherwise reducing sediments. This affects in particular the chemistry of iron and manganes, whose soluble reduced forms diffuse toward these interfaces, where they are precipitated as insoluble iron and manganese oxides. Thin reddish brown deposits on plant roots have been observed on several floodtolerant plants and attributed to the oxidative capacity of roots (23). Iron oxyhydroxides have been identified in these deposits (24). Other metals are also associated with these iron plaques (25), reaching higher concentrations than those found in the bulk sediments (Fig. 9). In regions where salt marsh plants remain active most of the year, like the Tagus estuary salt marshes (26), much thicker iron oxide coatings around the roots have been observed (25,27). Iron precipitates in tubular structures (Fig. 10), forming tubular concretions with a few millimeter size, so-called rhizoconcretions. These concretions are formed over a period of months (8) and contain more than 10% iron, doubling the concentrations of the bulk sediments (25). Rhizoconcretion formation thus contributes to salt marshes acting as natural sinks for trace elements from anthropogenic origin. 6

METAL ACCUMULATION IN ROOTS AND TRANSLOCATION

Besides sediments, salt marsh plants are also important temporary sinks for metals. Transition elements are accumulated in vegetation during the growth season (28). Roots, stems, and leaves of the salt marsh plants contain distinct levels of Zn, Pb, Cu, and Cd (Fig. 11). Zinc was the most abundant metal in all analyzed parts of the plants. The highest metal concentrations were recorded in the root system, and only small fractions were found in the above-ground parts of the plant. The accumulation partition was more pronounced for Pb and Cd, since these metals were more efficiently stored in the roots, than for Zn and Cu, which are translocated to the upper parts. The metal concentrations varied with the plant species and with the salt marsh. The results obtained in an annual study of the Tagus salt marsh indicate that Halimione portulacoides accumulates more Zn, Pb, Cu, and Cd than Spartina maritima (Fig. 11), and metal concentrations in their roots and sediments varied seasonally. The interaction between plant roots and sediments in salt marshes increases during the growing season and drops in winter, when plants decrease their activities (9,26). The temperature and salinity of the upper layer of Tagus salt marsh sediments rise in summer, to 30°C and 50°C, respectively. However, temperature is always positive in winter and salinity does not decrease drastically (27). Consequently, the salt marsh plants in the Tagus estuary remain active during the almost entire year. Sediments around Spartina maritima root were more acid and reductive (Fig. 12) than those around Halimione portulacoides, which indicates a stronger influence of that plant on sediments (9). However, pH and redox potential values did not show a pro-

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FIG. 11 Mean concentrations of Zn, Pb, Cu (µmol g⫺1 dry weight) and Cd (nmol g⫺1 dry weight) in roots ( ), stems (■), and leaves (䊐) of Spartina maritima and Halimione portulacoides (n ⫽ 8) between July 1991 and July 1992.

nounced seasonal variation, as would occur if plants stopped interacting with the sediment chemistry in winter. Daily inundation of the salt marshes by tide may also alter the sediment pore water chemistry, but its effect should be considered minor because conditions in the different plant areas vary. However, root biomass increased substantially between March and August/September, and the metal concentrations in both plants showed the same pattern: lower values in winter

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FIG. 12 Mean and standard deviation of the organic matter content (LOI), pH, and Eh (July 1991 to July 1992) of the sediments colonized by Spartina maritima (ⱓ) and Halimione portulacoides (䉬), n ⫽ 8.

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and an increase with the biomass. This means that growing process induces an increase of metal uptake by the roots. The root system is clearly the major pool of Zn, Pb, Cu, and Cd for Spartina maritima and Halimione portulacoides, since metal concentrations were higher in roots than in stems and leaves. These results are in agreement with the efficient storage of metals in roots that was observed for several salt marsh plants (28). The progressive increase of biomass from 3 to 8 kg m⫺2 (Fig. 13) implies that a considerable amount of metals was taken up by the roots: In autumn/winter part of the root system dies and metals returned to the sediments. These exchanges can only be registered in the bulk sediments when root biomass is abundant, as in the Tagus where the ratio of root to sediment reached 1:4. The increase of Zn, Pb, and Cu concentrations in the sediment that was registered in January, when levels in the roots and biomass were lower, corresponded to the transfer of metals from the plant to the sediments. In the spring, metal levels decreased in the sediment between roots (Fig. 14), as result of an inverse transportation of metals. This pattern was observed for the two studied plants (28). Cadmium showed a different seasonal distribution of the other three analyzed metals, suggesting that root-sediment exchange is more complicated. In spite of the high root density in sediments it is still a simplistic approach to consider that exchanges of metals involve roots and sediments exclusively. A mass balance of metals exchanged between sediment and roots was calculated for each 2 months for 1 year. Changes of metal concentrations in sediments were,

FIG. 13 Root biomass (kg m⫺2) of Spartina maritima (■) and Halimione portulacoides ( ), n ⫽ 3.

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FIG. 14 Mean and standard deviation of concentrations of Zn, Pb, Cu (µmol g⫺1 dry weight), and Cd (nmol g⫺1 dry weight) in sediments between the roots of Spartina maritima ( ) and Halimione portulacoides ( ) (n ⫽ 8) between July 1991 to July 1992.

in general, more accentuated than in roots. Processes other than uptake by the plants should occur in the sediments, namely, oxidation of insoluble sulfides (13) and formation of insoluble oxides (6) in microenvironments. These processes may interfere with the trace metal mobility (8,30). Only in winter were the Pb variations in sediments and roots comparable, suggesting a simpler transfer be-

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tween the two compartments. The uptake process by the roots involves the soluble metals in pore water, and their concentrations are influenced by several factors, namely, tidally induced advection, diffusion, and reactivity in the rhizosphere. In addition, the variation of mobile metals in sediments and roots may be poorly expressed by the total metal concentration. These geochemical processes and physical mechanisms all contribute to complicate a simple metal exchange between roots and sediments. Apart from the complexity of those exchanges, it is clear that Halimione portulacoides acumulates more metals than in Spartina maritima, and that plant appears to be a better temporary sink for metals than Spartina maritima. Our results are in agreement with a general rule that states that dicotyledons (Halimione portulacoides) have the capacity for higher accumulation of metals that monocotyledons (Spartina maritima) (31). 7

ROLE OF HALOPHYTES ON METAL CYCLING IN SALT MARSHES: A PROPOSED MODEL

The results obtained in these investigations allow the proposal of a simplified model for the retention of contaminants in the sediments and their interaction with the plants. A schematic representation is shown in Fig. 15. The steps of metal dynamics in salt marshes include the following: 1. Import of metals from the estuary in dissolved forms and associated with suspended sediments and biogenic particles transported by the tide. Higher enrichment in surface sediments occurs in the proximity of anthropogenic sources, highlighting the contribution of the salt marshes to remove anthropogenic metals from the contaminated estuarine waters. 2. Metal remobilization and subsequent alterations on metal fractionation in the sediments due to diagenetic reactions. The high quantity of sedimentary organic matter favors the internal metal redistribution. 3. Take-up of metals from contaminated sediments during the growth season and accumulation in the plant tissues. Basically through their subterranean components vascular plants in salt marshes act as temporary ‘‘sinks’’ for heavy metals. 4. Releasing oxygen to the rhizosphere, vascular plants alter the biogeochemistry of the sediments and modify dramatically the rooting sediments. Results from this interaction include modification of metal speciation around roots and buildup of iron-rich rhizoconcretions in which trace elements are sorbed. Accumulation of heavy metals in these structures may greatly exceed concentrations found in nearby sediments.

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FIG. 15 Schematic representation of factors affecting uptake of heavy metals. The model means illustrate that bioavailability of the elements is one of many factors determining uptake by plants and that the influence of root biochemistry is a key factor.

8

SUMMARY

Salt marsh vegetation influences the dynamics of the estuarine ecosystem and retains efficiently anthropogenic metals discharged to the system. Basically metal cycling in salt marshes includes retention of suspended particles and associated anthropogenic metals transported by the tides; taking up of metals from contaminated sediments during the growth season; accumulation in the plant tissues,

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mostly within the root system. In this way, vascular plants act as temporary ‘‘sinks’’ for heavy metals. Roots also act as an important vector to the incorporation of metals in salt marsh sediments. By releasing oxygen to the rhizosphere, vascular plants influence the biogeochemistry of the sediments and greatly modify the sediment characteristics of salt marshes. These interactions cause changes of metal speciation in the rhizosphere and buildup of tubular concretions around the roots rich in iron hydroxides. These concretions accumulate metals in concentrations higher than those found in nearby sediments, contributing to salt marshes that act as natural sinks for these elements. In unstable sediment environments such as the Tagus salt marshes, several processes change the availability of the metals to the plants. However, changes induced by the presence of roots are most prominent. Sediments around the roots are richer in organic matter and have a lower pH and higher redox potential than unvegetated bulk sediments. When compared with sediments not colonized by vascular plants, the metals in rooting sediments have lower concentrations of mobile chemical forms and, therefore, lower percentages of metals available to plants. The accumulation of metals in sediments and vegetation, together with the root-sediment interactions, all contribute to an overall reduction of metal bioavailability to plants. This suggests that salt marshes in the Tagus estuary may be important areas to help reduce the environmental contamination caused by the industrial and urban zones. Further investigation of the biogeochemistry of sediments and root-sediment interactions is needed to yield a better understanding of the reduction of metallic pollution in estuarine environments. REFERENCES 1. P Leendertse. Impact of nutrients and heavy metals on the salt marsh vegetation in the Wadden Sea. PhD thesis, Vrije Universiteit, Amsterdam, 1995. 2. ML Otte. Heavy metals and arsenic in vegetation of salt marshes and foodplains. PhD thesis, Vrije Universiteit, Amsterdam, 1991. 3. C Vale. Temporal variations of particulate metals in the Tagus river estuary. Sci Tot Environ 97/98:137–154, 1990. 4. P Adam. Salt Marsh Ecology. New York: Cambridge University Press, 1990, p. 445. 5. W Stephen, DE Seneca, W Woodhouse. Tidal salt marsh restoration. Aquatic Bot 32:1–22, 1988. 6. BL Haines, EL Dun. Coastal marshes. In: BF Chabot, HA Mooney, eds. Physiological Ecology of North American Plant Communities. London: Chapman and Hall, 1985, pp. 323–347. 7. I Cac¸ador, C Vale, FM Catarino. Accumulation of Zn, Pb, Cu, Cr and Ni in sediments between roots of the Tagus estuary salt marshes, Portugal. Est Coast Shelf Sci 42: 393–403, 1996. 8. B Sundby, C Vale, I Cac¸ador, F Catarino, MJ Madureira, M Caetano. Metal-rich

Retention of Heavy Metals in Salt Marshes

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concritions on the roots of salt-marsh plants. Mechanism and rate of formation. Limnol Oceanogr 43(2):245–252, 1998. ML Otte, J Rozema, L Koster, MS Haarssma, RA Broekman. Iron plaque on roots of Aster tripolium L.: interaction with zinc uptake. New Phytol 111:309–317, 1989. O Oenema, R Steneker, J Reynders. The soil environment of the intertidal area in the Westerschelde. Hydrobiolo Bull 22:21–30, 1988. RA Orson, RL Simpson, RE Good. A mechanism for the accumulation and retention of heavy metals in tidal freshwater marshes of the upper Delaware River estuary. Est Coast Shelf Sci 34:171–186, 1992. I Cac¸ador, C Vale, FM Catarino. The influence of plants on concentration and fractionation of Zn, Pb, and Cu in salt marsh sediments (Tagus Estuary, Portugal). J Ecosyst Health 5:193–198, 1996. MJ Madureira, C Vale, ML Simo˜es. Effect of plants on sulphur geochemistry in the Tagus salt-marsh sediments. Mar Chemistry 58:27–37, 1997. B Tinker, P Barraclough. Root-soil interactions. In: O Hutzinger, ed. Reactions and Processes, 2: Part D. Berlin: Springer-Verlag, 1988, pp. 154–171. JJ Albertsl, MT Price, M Kania. Metal concentrations in tissues of Spartina alterniflora (Loisel) and sediments of Georgia salt marshes. Est Coast Shelf Sci 30: 4–58, 1990. J Rozema, ML Otte, R Broekman, G Kamber, H Punte. The response of Spartina angelica to heavy metal pollution. In: AJ Gray, ed. Spartina angelica. London: HMSO, 1990, pp. 39–47. M Caetano, M Falca˜o, C Vale, MJ Bebbiano. Tidal flushing of ammonium, iron and manganese from inter-tidal sediment pore waters. Mar Chem 58:203–211, 1997. I Cac¸ador, C Vale, FM Catarino. Effects of plants on the accumulation of Zn, Pb, Cu and Cd in sediments of the Tagus estuary salt marshes, Portugal. In: J-P Vernet, ed. Environmental Contamination. Amsterdam: Elsevier, 1993, pp. 355–365. PM Chapman. The sediment quality triad approach to determining pollution-induced degradation. Sci Total Environ 97/98:815–825, 1990. I Cac¸ador, C Vale, F Catarino. Relationships between metal concentrations and organic matter content in the Tagus estuary salt marsh sediments halophyte uses in different climates, I. In: A Hamdy, et al. eds. Leiden: Backuys Publishers, 1999, pp. 103–110. A Piccolo. Reactivity of added humic substances towards plant available heavy metals in soils. Sci Tot Environ 81/82:607–614, 1989. BJ Alloway. Soil processes and the behavior of metals. In: BJ Alloway, ed. Heavy Metals in the Soils. Glasgow: John Wiley & Sons, 1990. IA Mendelssohn, M Postek. Elemental analysis of deposits on the roots of Spartina alterniflora Loisel. Am J Bot 69(6):904–912, 1982. AA Crowder, SM Macfie. Seasonal deposition of ferric plaque on roots of wetland plants. Can J Bot 64:2120–2124, 1986. C Vale, FM Catarino, C Cortesa˜o, MI Cac¸ador. Presence of metal-rich rhizoconcretions on the roots of Spartina maritima from the salt marshes of the Tagus estuary, Portugal. Sci Total Environ 97/98:617–626, 1990. FM Catarino, MI Cac¸ador. Produc¸a˜o de biomassa e estrate´gia de desenvolvimento

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27. 28.

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Cac¸ador and Vale em Spartina maritima e outros elementos da vegetac¸a˜o dos sapais do estua´rio do Tejo. Bol Soc Broteriana 2a se´r 54:384–403, 1981. I Cac¸ador. Acumulac¸a˜o e retenc¸a˜o de metais pesados nos sapais do Estua´rio do Tejo. Tese de Doutoramento, Universidade de Lisboa, 1994. I Cac¸ador, C Vale, FM Catarino. Seasonal variation of Zn, Pb, Cu and Cd concentrations in the root-sediment system of Spartina maritima and Halimione portulacoides from Tagus estuary salt marshes. Marine Environmental Research 49:279–290, 2000. WHO Ernst. Element allocation and (re)translocation in plants and its impact on representative sampling. In: H Lieth and B Markert, eds. Element concentration cadasters in ecosystems. Weinheim: VCH Verlagsgesellschaft, 1990, pp. 17–40. WG Beeftink, J Rozema. The nature and functioning of salt marshes. In: W Salomons, BI Bayne, EK Dursma, and U Forstner, eds. Pollution of the Northsea: An Assessment. Berlin: Springer-Verlag, 1988, pp. 59–87. WHO Ernst, W Mathys, J Salaske, P Janiesch. Aspekte von Schwermetallbelastungen in Westfalen. Abhandl. Landesmus. Naturk Westf 36(2):1–30, 1974.

5 Lichens Cristina Branquinho Universidade Atlaˆntica, Oeiras, Portugal

1 1.1

INTRODUCTION Understanding the Singularity of Lichens

Lichens are by definition symbiotic organisms composed of a fungal partner, the mycobiont, and one or more photosynthetic partners, the photobiont, which may be either a green alga or a cyanobacterium (1). The nature of the lichen symbiosis is still not totally clear: although most authors consider lichens to be a classic example of mutualism, in which all partners gain benefits from association, others suggest that lichens are an example of controlled parasitism (1). The degree of lichenization varies significantly from photobiont cells loosely attached to a fungus to a more complex organization that in no way resembles the bionts that form it (1). Lichens are a very successful symbiosis; they are found in almost all terrestrial habitats, including tropical, desert and polar, and in fresh and salt water habitats. Lichens are dominant life forms in 8% of the earth’s terrestrial surface (2). The mycobiont, a heterotrophic organism, benefits from the carbohydrates produced by the photobiont. In the case of nitrogen-fixing cyanolichens, the mycobiont also gains a nitrogen source. The fungus probably functions as a reservoir 117

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of inorganic nutrients for the photobiont and also creates a humid environment, enhancing water uptake, offers a physical support, and substantially reduces the light intensity for the photobiont. On the basis of their overall habitat, lichens are traditionally divided into three main morphological groups: crustose, foliose, and fruticose. Crustose lichens are tightly attached to the substrate with their lower surface and may not be removed from it without destruction. Foliose lichens are leaf-like, flat, and only partially attached to the substrate. Fruticose lichens always stand out from the surface of the substrate (Fig. 1). Their thallus lobes are hair-like, strap-shaped, or shrubby and the lobes may be flat or cylindrical. Lichens may also be classified according to the substrate they occupy: as endolithic if they grow inside rock; epilithic if they grow on rock surfaces; terricolous if they grow in soil; and epiphytic if they grow on trees. Lichens are extremely slow-growing and long-lived organisms whose growth rates mostly depend on climate conditions. Lichens of extreme climates, such as desert, polar, or alpine ecosystems, have only short periods in which full metabolic activity and growth can occur. Consequently, only very low cell turnover rates and minimal annual size increases are recorded (⬍1 mm/year). However, the majority of lichens in temperate or subtropical to tropical climates have annual radial growth in the range of millimeters to a few centimeters. Although

FIG. 1 The fruticose lichen Evernia prunastri.

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extreme ages of lichens are estimated in the range of millennia (3), the majority develop over years or decades. Estimates of lichen longevity should be interpreted critically because the metabolically active cells may be not as old as the thalli. 1.2

Chemical and Physical Properties of Metals

The position and classification of the chemical elements in the periodic system of the elements (PSE) do not indicate their acute or chronic toxicity to living organisms (4–6), since the PSE is based on purely physicochemical characteristics (6,7). Proposals have been made for a classification of metals more relevant to biological systems, based on the chemical behavior of the metals’ cations and especially on their ability to become part of a chemical complex (4,5,7–9). Nieboer and Richardson (4) divide elements into three groups: (a) class A elements showing preference for ligands containing oxygen, including all of the macronutrient cations of terrestrial plants (e.g., K, Mg, Ca), as well as some other alkaline and earth metals (e.g., Cr and Fe); (b) class B elements showing preference for ligands with nitrogen or sulfur (e.g., Ag, Au, and Hg); and (c) borderline elements, with intermediate characteristics between class A and class B (e.g., Cu, Co, Ni, and Pb). 1.3

Rationale of the Chapter

From the above the complexity and singular nature of these organisms are obvious. Moreover, there is a great lack of information relating to both the nature of the lichen symbiosis and the interactions between the bionts. Thus, lichens should always be regarded as complex, mostly unknown biological systems. The role of lichens as sensitive indicators and monitors of atmospheric pollution is of undeniable importance. Despite the great number of publications dealing with the interaction of lichens with heavy metals (10), a detailed knowledge of the physiology of metal absorption in these organisms, i.e., uptake, accumulation, retention, localization, release, tolerance, and toxicity of metals, is needed before lichens can be extensively and reliably used for monitoring of metal deposition patterns (11–14). In this chapter a critical review of the more pertinent aspects of the interaction between lichens and metals will be given with emphasis on effects of metals on lichen communities and on physiological performance; their tolerance mechanisms; sources, retention, and release of metals; uptake and location of metals; sources of variability during lichen sampling; and calibration of lichens against dust gauges. 2

LICHENS AS GOOD BIOINDICATORS AND BIOMONITORS OF METAL POLLUTION

Lichens are the most studied bioindicators and biomonitors of air pollution. Some lichens are extremely sensitive to atmospheric pollution and have been used as

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good bioindicators of such pollution for more than a century (15–17). The importance of this role is evidenced by the large number of publications on this topic (10). Lichens have also been used extensively to biomonitor heavy metal contamination (11,15,16,18). Bioindicators can provide information on qualitative variations in the environment, while biomonitors supply quantitative information. Both uses are based on organisms, communities, morphology, cellular structure, metabolism, and biochemical processes, etc. (19,20). The concentrations of given pollutants, measured within the organism, are used to reconstruct the spatial and temporal deposition patterns of the pollutants deposited at a location (18,19). 2.1

Lichens as Bioindicators

Epiphytic lichens have traditionally been used as bioindicators of atmospheric pollution because they show differential sensibilities to air pollution, i.e., the most sensitive lichens tend to disappear from polluted areas whereas the most tolerant species can be seen in areas with moderate pollution emissions (9,15,17,21,22). The first documented reports of the disappearance of lichens from city centers were made in the 1800s (23). Nylander (24) suggested that the absence of the lichens was due to air pollution. In 1970, Hawksworth and Rose (25) correlated the declines of lichens in urban areas with mean winter SO2 measurements and developed a semiquantitative scale for English forests relating the occurrence of approximately 50 lichen species. Since then almost every city in Europe and some in Canada and the United States have been mapped using lichen biodiversity data (19). Today the list of air pollutants related to the decline of lichens in urban and industrial areas is much longer, including hydrogen fluoride (HF), heavy metals, acid rain, radionuclides, oxidants and organic compounds, and it is not complete (23). 2.2

Lichens as Biomonitors

Lichens have been used extensively to biomonitor heavy metal contamination from anthropogenic sources: roads, mines, and industrial facilities in urban and rural locations (11,15,16,18). However, they have also been used to assess natural sources of metals, such as volcanoes (26). Many authors have pointed out that lichens have several of the characteristics required of the ideal biological monitor (11,15,16,18,19,27). They have a wide geographic distribution, occurring in rural, urban, and industrial areas, thus allowing comparison of pollutant concentrations from diverse regions. The morphology of lichens does not vary seasonally, so that pollutant accumulation can occur throughout the year. They are photosynthetically active throughout the year. Their ability to accumulate heavy metals has been widely demonstrated (11,28,29). Lichens are poikilohydric organisms, lacking a developed cuticle, which have very limited control of the uptake and loss of water and solutes from atmospheric deposition. Furthermore, they depend

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mostly on atmospheric deposition for nutrition because their ‘‘root structures’’ do not function as in higher plants. They lack mechanisms for active absorption from the substrate. On the other hand, in contrast to higher plants, they have no excretion mechanisms. In situ lichens are slow-growing, long-lived organisms, which facilitates their use as long-term integrators of atmospheric deposition. Lichens are easily handled and transplanted, thus permitting biomonitoring for defined periods of time and the assessment of short-term atmospheric deposition. Both bioindicators and biomonitors may be considered as organisms or communities of organisms that react to variations in environmental conditions with a change in their ‘‘signs of activity’’ (20). These changes may manifest themselves: 1. In the structure or dynamics of the population or community of organisms; 2. In changes in the organisms’ functional response that may lead to changes in vitality; 3. In concentrations of elements that influence the organism. 3

MAPPING LICHEN BIODIVERSITY TO ASSESS METAL POLLUTION

Lichens may be used as bioindicators whenever air pollution causes compositional changes in their communities. The principal outputs in this area are biodiversity maps based on all lichen flora and on the use of several indices of atmospheric purity (IAPs). Over the last 30 years, numerous studies have assessed the effect of air pollution on lichens, which supposedly reflect the local ‘‘air quality.’’ This is usually accomplished by mapping of lichen communities. Lichens have been used as bioindicators to evaluate pollution in many urban and industrial areas; more than 50 cities around the world have already been mapped (19). Several regional and national studies have been performed in England, the Netherlands, Finland, Denmark, Sweden, Slovenia, Estonia, southern California, northern Italy, etc. (19). The importance of lichen biodiversity data was recently demonstrated by Cislaghi and Nimis (30), who compared such data with mortality maps of a large part of northeastern Italy. The results strongly supported a relationship between air pollution measured by the lichen biodiversity data and lung cancer. One of the most common approaches to mapping lichen biodiversity is the use of IAPs. The IAP originally developed by LeBlanc and De Sloover (31) is based on the presence of lichens, on their frequency, and on an ecological index for the species. This toxitolerance factor for each species is determined on a somewhat subjective basis (17,32). Furthermore, there is evidence that the tolerance of a given species to air pollution may differ according to the general climatic conditions (17). Thus, even in the rare cases where a toxitolerance factor

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for calculating the IAP is based on experimental data, its extrapolation to another area, with different climatic conditions, may not be accurate and cannot be recommended (17). Both the lichen frequency and the number of lichen species have been shown to be less subjective parameters that are correlated with changes in atmospheric pollution (14,19,32). The declines in lichen communities found in the majority of biodiversity studies were mainly due to gaseous pollution (mainly HF and SO2) (19). Whenever there is a mixture of both gaseous and metal pollution, most authors have difficulty in distinguishing the relative contribution of metal pollution to lichen disappearance or to the injury symptoms observed under field conditions. The distribution of epiphytic lichen species has been studied in the area of a copper smelter in Canada and around a zinc smelter in the United States (33,34). These studies suggested that the lichen distribution was influenced more by SO2 emissions than by metal pollution. Few studies have reported the effects on lichen flora of metal pollution without gaseous emissions. Folkeson and Andersson-Bringmark (35) studied the lichens in the coniferous forest surrounding the brass mills at Gusum, southeastern Sweden, which is heavily polluted with Cu and Zn but not with SO2. The author concluded that the deterioration of the lichen vegetation was due to Cu and Zn. The effects on the disappearance of some species of the genus Cladonia due to metal pollution were clearly observed (35). Loppi et al. (36) showed the results of a retrospective study performed in the Chianti region in central Italy using lichen biodiversity data and concentrations of elements accumulated in the lichen Parmelia caperata. The results of sampling in 1980 and 1996 showed that the general decrease in concentrations of metals with time was associated with an increase in the number and frequency of lichen species (36). Herzig et al (32) combined lichen biodiversity data with multielemental analysis in lichens and found that many elements correlated well with the IAP18 (based on the frequency of more than 40 lichen species). They suggested that biodiversity and multielemental analysis are complementary, enabling detailed statements on biological effect and total air pollution in general and on single pollutants in particular. Another study evaluated the impact on lichen flora of an underground copper mine located in southern Portugal (14,37). The mine processes sulfide ore to produce Cu concentrates that are transported by rail and shipped to smelters overseas. The main source of atmospheric pollution at the mine site is fugitive dust emissions from the stockpiles and waste heaps. Since the mine is located in a rural area distant from any other industry or major city, there are no other significant local sources of air pollution (14,37). This study used lichen biodiversity, measured by the parameter number of lichen species, to identify the zones of the mine subject to long-term emissions (Fig. 2). In this case study, the fruticose

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FIG. 2 Estimated spatial distribution of the total number of epiphytic lichens in the surroundings of a copper mine obtained by kriging of data based on 19 sampling points (⫹) during 1993. The center of the mine is represented by 䊉, the Cu-concentrate stockpile by 䉱, and the waste heaps by ■. The studied area was divided into classes with different numbers of lichen species: 䊐 ⬍10, ■ 10–18, and ■ ⬎18 species. (From Ref. 37.)

lichens appeared to be the most sensitive to Cu pollution, followed first by the foliose and then by the crustose forms (14). One of the difficulties with mapping lichen biodiversity is that it requires specific training in lichenology. Recorders must be competent at identifying and distinguishing a wide range of lichens in different stages of development. The use of lichen data on biodiversity maps to evaluate the impact of atmospheric pollution should be accompanied by the concentration of elements accumulated in the lichen thallus to avoid incorrect interpretations (14,32). Lichens are also bioindicators of several other environmental factors such as climate, so that changes in the composition of a lichen community may not be solely due to changes in pollution levels. Once the primary pollutants that cause damage to lichen communities have been identified, the levels of the pollutants that cause those changes should be evaluated (38).

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The main advantage of lichen biodiversity mapping is that it combines a high number of sampling points with relatively low costs, enhancing the data quality. As a first approach, such studies allow the easy identification of the highrisk areas to be further examined in future studies. An example is the optimization of the location of chemical and physical samplers whose numbers are limited due to their high costs. Biodiversity mapping studies are also important as effective early warning systems to detect environmental changes. 4

PHYSIOLOGICAL RESPONSES OF LICHENS TO METALS

The main topics considered in physiological studies are morphological changes, variations in vitality, and changes in functional response. Under field conditions it is difficult to determine the effect of heavy metals on the physiology of lichens because their deposition is often linked with acid rain (SO2) or the presence of O3, nitric oxides (NOx), peroxyacetyle nitrate (PAN), HF, or other anthropogenic gases (39). Most studies on the effects of metals on lichens were performed under controlled conditions, and only a few have been confirmed under field conditions (37). 4.1

Effect on Cell Membranes

The cell membrane is an obvious initial site for toxic metal action, and the loss of ions, particularly K, from the cell interior in lichens has been widely used as an indication of membrane integrity (12–14,37,40–44). This is based on the fact that most of the K is located intracellularly. A proportion of Mg is also readily available within the cell, and this cation may also be used as an indicator of severe membrane damage (12–14,37,45,46). The recovery of soluble elements from incubation media or washing solutions, measured by conductivity, has been used to indicate cellular damage: by metals in the laboratory or by air pollution in the field (39,47–50). Estimation of K loss into such washing solutions might overestimate ‘‘leaching from the cell’’ if there is a source of this element in the environment. At sites with very high levels of dry deposition, K deposited on the surface or bound to the wall of the lichen may reach 50% of the total K concentration (Fig. 3). Thus, the conductivity may not be always related to K loss from inside the cell (51). The sequential elution procedure was shown to be valuable for distinguishing between elements acquired from the environment and those physiologically released (see Sec. 6). Some authors wash the lichen a few seconds before the conductivity measurements as described by Garty et al. (39), but care must be taken because a small proportion of K might still be bound to the wall and be displaced by other cations with a stronger affinity for those binding sites. The ion loss is a measure of the physiological state of the mycobiont, which composes approximately 90% of the lichen dry weight (12,52). Tarhanen et al.

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FIG. 3 Concentration of Ca (A and B), Mg (C and D), and K (E and F) in extracellular soluble, intracellular soluble, and particulate fractions in (䊐) marginal zones and (■) central zones of the lichen Xanthoria parietina collected near a cement mill (A, C, and E) and at Lisbon (B, D, and F). Columns are means and bars are the standard deviations of five replicates. For the same element, site and location values with different letters are significantly different for P ⬍ 0.05 using a one-way ANOVA. (From Ref. 51.)

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(50) supported the hypothesis that ergosterol is a fungal-specific component and there is a significant correlation between ergosterol content and the K leakage. It is the dominant sterol in lichenized fungi and a principal constituent of the fungal plasma membrane; its concentration is indicative of the metabolically active fungal biomass. Changes in sterol composition can alter the physical properties of membranes and fatty acid component of lipids, and can collectively influence enzyme activity and ion permeability. The effect of metals on lichens depends primarily on the species and on the concentration applied. Under controlled conditions, the cations Hg, Tl, and Ag have been shown to cause severe membrane damage in lichens (29,40). Levels of Ni, Mn, Co, and Cd only cause membrane damage under unrealistic conditions, such as low pH and very high concentrations (29,40). However, one of the elements that undoubtedly causes severe membrane damage in lichens is Cu (13,14,29,37,40,42,44,50,53). This element also causes membrane damage in higher plants (8,54), fungi (5), and hepatics (55). The effects of supplied Cu on the release of physiological elements from the lichens Usnea spp. and Ramalina fastigiata were investigated to establish the extent of Cu-induced membrane damage (13). The study hypothesized that Usnea spp. would be more sensitive than R. fastigiata to Cu because a lichen biodiversity study around a Cu mine in southern Portugal had found the latter to have a wider distribution than the former (14,37,56). To establish the relative sensitivity of the lichens to Cu uptake, intracellular Cu concentrations were correlated with alterations in membrane integrity (Fig. 4). It was shown that Usnea spp. was more sensitive to Cu uptake than R. fastigiata, with greater K loss occurring at lower supplied Cu concentrations (Fig. 4). In both lichens losses of intracellular K (Fig. 4 and Table 1) appeared to occur as a result of intracellular Cu uptake. The authors found no losses of intracellular Mg (Fig. 4) and suggested that Cu induces specific K loss in lichens rather than nonselective membrane damage (46), arguing that Mg is also lost when generalized membrane damage occurs, as with desiccated samples (13,41). In an experiment performed under field conditions, Tarhanen et al. (50) also reported membrane damage, measured through K leakage and ergosterol concentration, when the metal content of Bryoria fuscescens thalli exceeded 6.3 µmol g⫺1 Cu and 1.7 µmol g⫺1 Ni. The effect of Pb on membrane integrity is controversial: some authors have reported slight membrane damage at very high concentrations (40,42), whereas others have not (12,14). Nieboer et al. (42) reported a slight K efflux due to Pb uptake on Umbilicaria muhlenbergii. However, these authors did not discriminate between extra- and intracellular K, and some of the K efflux may have derived from extracellular sites. Branquinho et al. (12) investigated the effect of extraand intracellular Pb concentration on membrane integrity of several epiphytic lichen species with different photobiont associations (cyanobiont, phycobiont, and phycobiont with cyanobacterial cephalodia) and thallus growth forms (foli-

FIG. 4 Response on incubation for 2 h in different Cu concentrations of intracellular mean concentrations of Cu (䊉), K (䉬), Mg (■) and Ca (䉱) in Ramalina fastigiata (open symbols) and Usnea spp. (closed symbols). The control sample (c) was incubated in DDW for 2 h. Symbols are means and bars are the standard deviations of five replicates. (From Ref. 13.)

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TABLE 1 Linear Correlation Coefficients (Pearson r) of Intracellular Cu Concentrations with Changes in the Distribution of Intracellular K, Mg, and Ca Concentrations in Ramalina fastigiata and Usnea sppa Lichen species R. fastigiata Usnea spp.

K

Mg

Ca

⫺0.82** ⫺0.57*

⫺0.47 0.24

⫺0.13 ⫺0.09

a Cation measurements were performed after incubation for 2 h in solutions of different Cu concentrations. Significant linear correlations are marked * for 0.1 ⬎ P ⬎ 0.05 and ** P ⱕ 0.05, n ⫽ 9. Source: Ref. 14.

ose, fruticose) (Fig. 5). Increasing Pb concentration was not observed to alter lichen membrane permeability, as measured by the loss of intracellular K and Mg, in the species studied (Fig. 5 and Table 2). Also, no differences were observed in lichen membrane permeability with time after Pb incubation (120 h) (Fig. 6), and so no effect of Pb uptake was observed on the mycobiont (12). Under field conditions the concentrations of several elements—B, Al, Cr, Fe, Si, Ti, Zn (49,57) and Mg (58)—have been correlated with membrane damage as measured by electrical conductivity. However, this was not subsequently confirmed in the laboratory. This is important since the most dominant pollutant in several of these studies is SO2 (39,49,53,57), which also causes severe membrane damage, and it is difficult to isolate the effect of metals with certainty. 4.2

Effect on Photosynthetic Apparatus

Chlorophyll in lichens is usually determined by the method proposed by Ronen and Galun (59), who showed that the ratio of optical density between 435 nm and 415 nm is a reliable parameter for estimation of chlorophyll degradation in DMSO pigment extracts (48,53,60). Several authors indicate that metals increase the chlorophyll degradation (53,60). The metals responsible for significant increases in chlorophyll degradation in both the field and the laboratory are Pb, Fe, Zn, Mn and Cu (49,53,60). It is also suggested that the presence of K ions protects the photobiont chlorophyll against degradation and possibly stimulates chlorophyll synthesis (53). No effect was observed in chlorophyll degradation with increasing concentrations of Ca, Cl, Zn, St, and P. A displacement and a decrease in the absorption spectrum of chlorophyll with Cu, Hg, and Ag under controlled conditions was observed (40). Barto´k et al. (61) found that when transplanted for 150 days to an industrial area of Romania (cellulose manufactures and

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FIG. 5 Intracellular (A, B, C) K (A), Mg (B), and Ca (C) concentrations (µmol g⫺1) in lichens, immediately after 2 h incubation in different supplied Pb concentrations (presented in a logarithmic scale). For each species and cation, the control sample (c) was incubated in DDW for 2 h. The symbols show the average and the bars the standard deviation of three replicates for each lichen species: Peltigera canina (䉬), Lobaria pulmonaria (■), Parmelia caperata (䉱) and Ramalina farinacea (䊉). (From Ref. 12.)

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TABLE 2 Correlation Coefficients (Pearson r) of Intracellular Pb Concentrations with the Concentration of Intracellular K, Mg, and Ca in Four Lichen Species Species Peltigera canina Lobaria pulmonaria Parmelia caperata Ramalina farinacea

K

Mg

Ca

⫺0.44 ⫺0.58 0.25 ⫺0.86

0.92* ⫺0.76 0.88* 0.84

⫺0.03 ⫺0.46 ⫺0.42 ⫺0.65

Significant linear correlations are marked for * P ⬍ 0.1, n ⫽ 5. Source: Ref. 14.

refractory material plants), the lichen Xanthoria parietina underwent a significant decrease in the content of pigments, particularly chlorophyll a, which significantly correlated with accumulation of Mg, Cr, Fe, and Cd. Photosynthesis is more sensitive to Cu, Cd, and Zn in cyanobiont than in phycobiont lichens (62). This sensitivity appears to be related to the prokaryotic nature of the cyanobacteria; the algal component is thus more sensitive to metals. For Puckett (40), Ag and Hg followed by Cu were the metals which cause the greater reductions in 14C fixation, while for Richardson et al. (63), only Cu caused substantial reduced 14C fixation in Umbilicaria muhlenbergii, which was not affected by uptake of Sr, Mg, Ca, Ni, Zn, and Pb. Chlorophyll fluorescence has been used to study the effect of contaminants on the electron transport chain associated with photosystem II (PSII) in lichens (12–14,37,64–67). The ratio of variable chlorophyll fluorescence to maximal fluorescence (Fv /Fm) has been widely used as a parameter to assess the state of the photosynthetic apparatus and to reflect the efficiency of the primary photochemical reactions in PSII in higher plants (68–70) and in lichens (71,72). The effects of pollutants on lichens showed significant reductions in Fv /Fm with no changes in other photosynthetic parameters, showing that chlorophyll fluorescence was a more sensitive indicator of damage (64,66). The effect of extra- and intracellular Pb concentration on chlorophyll fluorescence in several epiphytic lichen species with different photobiont associations (cyanobiont, phycobiont, and phycobiont with cyanobacterial cephalodia) and thallus growth forms (foliose, fruticose) has been investigated (12). It was shown that Pb uptake resulted in a decrease in PSII photochemical reactions, particularly in cyanobiont lichens (Fig. 7) (12). A significant reduction of Fv /Fm in L. pulmonaria and P. caperata only occurred 48 h after Pb incubation, whereas in P. canina the decrease was observed immediately (Fig. 7). Although Pb is currently considered to be a metal with very low mobility (73), transfer of some wallbound Pb to the cell interior, as shown for other elements (Zn, Cu, and Cd) (43),

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FIG. 6 Effect of storage after Pb incubation in intracellular K (A) and Mg (B) concentrations, in different lichen species: Peltigera canina, Lobaria pulmonaria, Parmelia caperata, and Ramalina farinacea, submitted to three different treatments: (䊐) control samples assessed immediately after incubation in H2O; ( ) control samples assessed 120 h after incubation in H2O; and ( ) samples assessed 120 h after incubation in 4.8 mM PbNO3. Columns are averages and bars are standard deviations of three replicates. For each species and element, treatments with the same letter are not significantly different using a Tukey test for P ⬍ 0.05. (From Ref. 12.)

may have occurred in L. pulmonaria and P. caperata. The substantial decrease in Fv /Fm observed for L. pulmonaria might be related to a higher sensitivity of Dictyochloropsis algae to Pb uptake than Trebouxia (P. caperata and R. farinacea). Thus, the studied effects of Pb in lichens were shown to be dependent only on the nature of the photobiont. Our results showed that lichen PSIIphotochemical reactions were more sensitive to Pb than indicators of cell membrane damage (Figs. 4–7) (62,74).

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FIG. 7 Changes in average chlorophyll a fluorescence parameter (Fv /Fm) with time after Pb incubation (h) in different lichen species: Peltigera canina, Lobaria pulmonaria, Parmelia caperata, and Ramalina farinacea. Prior to the measurements the samples were incubated in different Pb concentrations for 2 h (mM): control (䊐), 0.005 (䉭), 0.05 (䊊), 0.48 (■), and 4.83 (䊉). (From Ref. 12.)

The effect of the Cu on lichen PSII photochemical reactions has been investigated under both field and laboratory conditions (37). There was a significant correlation between the decrease in Fv /Fm and the increase in intracellular Cu concentrations under field conditions (Fig. 8). In fact, at intracellular Cu concentrations above about 2.0 µmol g⫺1, the photochemical reactions of the lichen R. fastigiata were completely inhibited at the PSII level (Fig. 8). These results are in agreement with controlled Cu uptake experiments performed on R. fastigiata

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FIG. 8 Variation of chlorophyll fluorescence parameter (Fv /Fm) with intracellular Cu concentration in the lichen Ramalina fastigiata collected at different distances from the mine (r ⫽ ⫺0.88; P ⬍ 0.05; n ⫽ 14). Symbols represent the means and bars the standard deviations of five replicates. (From Ref. 37.)

(Fig. 9). The Fv /Fm values of R. fastigiata samples subjected to Cu addition under controlled conditions were zero for supplied Cu concentrations above 1.6 mM (with intracellular Cu concentrations between 3.7 and 4.3 µmol g⫺1) (Fig. 9). A highly significant (P ⬍ 0.01) linear correlation (correlation coefficient r ⫽ ⫺0.90, n ⫽ 10) was found between the Fv /Fm values and the intracellular Cu concentration. It was concluded that total inhibition of PSII photochemical reactions occurred in R. fastigiata under field and controlled conditions when intracellular Cu concentrations exceeded a threshold of about 2.0 µmol g⫺1. No samples of this species have been found under field conditions beyond the Cu threshold (37). These results are in agreement with the Cu threshold concentrations obtained for other lichen species by Folkeson and Andersson-Bringmark (35), which found the first signs of reduction in epiphytic lichen cover on pine trunks and twigs when concentrations of Cu in Hypogyminia physodes exceeded 2.0 µmol g⫺1. Thus, the fluorescence parameter Fv /Fm proved to be a good indicator of the survival capacity of R. fastigiata under field conditions and a useful parameter for determination of the sensitivity of lichens (photobiont) to Cu pollution. It would be useful to determine whether other lichen species have the same Cu threshold.

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FIG. 9 (䊉) Response of chlorophyll fluorescence (ratio Fv /Fm) in the lichens Ramalina fastigiata after incubation for 2 h in different Cu concentrations. Symbols are the means and bars the standard deviations of three replicates. (䊊) Response on incubation for 2 h in different Cu concentrations of intracellular mean concentrations of Cu in Ramalina fastigiata. Symbols are means and bars are the standard deviations of five replicates. The control sample (C) was incubated in deionized water for 2 h. (From Ref. 37.)

4.3

Other Effects

Current information on ultrastructural effects due to metals alone is insufficient (75). Copper and nickel (Ni) treatments applied under field conditions induced ultrastructural changes in algal and fungal cells of the lichen Bryoria fuscescens (75). The response depended on the nature of the metal, its concentration, and climate. The photobiont was the most sensitive component, and the main ultrastructural changes in alga cells were located in the chloroplasts and mitochondria (75). The photobiont developed ultrastructural injuries very quickly when thallus metal concentrations exceeded 0.9–1.2 µmol g⫺1 Cu and 0.12–34 µmol g⫺1 Ni in combination with acidity, or more than 20 µmol g⫺1 Ni in the absence of acidity. Recently, Garty and co-workers (39,58) used the spectral reflectance response of lichens as an indicator of the physiological status of the alga cells, stating that spectral reflectance correlates empirically with photosynthetic activity, fractional vegetation cover, green leaf biomass or leaf area, primary produc-

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tivity, and carbon in standing biomass. The authors demonstrated that remote sensing methods effectively detect lichens under stress induced by air pollutants such as heavy metals and SO2 (39,58). The normalized difference vegetation index (NDVI) was positively correlated with K concentration in the lichen thallus and negatively correlated with the concentration of Mn, Ni, Pb, and S. Other authors have used the production of ethylene as an indicator of metallic stress in lichens (58,76). Lichens display an increase in the ethylene production when exposed to air polluted with heavy metals (11,58,76,77); this increase reduces the lichens’ growth rate. Endogenous auxins in lichens 1 year after transplantation to an air-polluted urban site were lower than at the control site (77). Barto´k et al. (61) found that when transplanted for 150 days to an industrial area of Romania (cellulose manufactures and refractory material plants), the lichen Xanthoria parietina showed a decrease in the respiration rate (measured through dehydrogenasic activity) with signs of total exhaustion. The respiration rate was correlated with the distance from the pollution source. The lichens also significantly accumulated Mg, Cr, Fe, and Cd. Glenn et al. (78) studied the influence of vehicular traffic on lichens growing in the natural reserve of Montseny, Catalonia, Spain, and found signs of damage to the lichens. These were mainly due to arthropod feeding and fungal parasites, which were substantially elevated on roadsides (14–33%), in comparison with a control site (10%). They also found elevated levels of Pb, Zn, and Cu in large thalli near the roadsides. 4.4

Comment

In general, the primary effects of heavy metals are dependent on the nature of the photobiont, the symbiotic partner most sensitive to heavy metals. The cyanobiont lichens have been shown to be more sensitive than the phycobiont ones. Of the most studied elements, Cu and Hg have been shown to be most toxic to lichens. The applicability of the chlorophyll fluorescence Fv /Fm ratio as a monitor of metal pollution effects on lichens has been demonstrated. The method has the advantage of being nondestructive and noninvasive, facilitating repeated measurements. One of the main problems of using lichens as bioindicators of atmospheric metal pollution is the difficulty of demonstrating the extent to which morphological and physiological changes observed under field conditions are due to metal pollution alone. Lichens are also quite sensitive to variations in other abiotic factors, such as moisture, light, temperature, and pH. When evaluating lichen biodiversity or the physiological effects of pollutants under field conditions, care must be taken to determine whether other abiotic factors have also changed and to confirm the effect of the pollutants under controlled conditions. On the other hand, ecophysiologists should also verify pollution conditions before attributing

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the decrease in lichen performance to changes in climate, microclimate, pH, forest degradation, and so forth. The various authors are not always in agreement as to the physiological effects of metals on lichens. It is suggested that future studies consider the metal concentrations that actually enter the cells to justify stated physiological impacts; otherwise it will be difficult to define threshold values for the organisms. 5

SOURCES, RETENTION, AND RELEASE OF METALS IN LICHENS

The concentration of elements in lichens is the result of the equilibrium between uptake and release from and into the surrounding environment. One of the drawbacks of using biological systems for monitoring is that the mechanisms of element uptake and retention by biomonitors are still not sufficiently known (27). 5.1

Wet Deposition

Elements in the atmosphere are scavenged from the air and intercepted by lichens through precipitation and occult precipitation (fog and dew). The latter is very important to lichens, and concentrations of nutrients and contaminants may be substantially higher in fog and dew than in rainfall because more dilution occurs in the formation of rain (79). Few studies have considered the effect of occult precipitation on the concentrations of elements in lichens. Concerning the rainfall process, there is no consensus regarding the contribution to element concentration in lichens. The effect of rainfall on Cu and Fe concentrations in lichens transplanted for 3 months to a Cu mine site in the south of Portugal was studied over 3 years (Table 3). Total Cu and Fe concentrations increased in R. fastigiata and Usnea spp. with increasing rainfall volume, particularly if the rainfall had occurred during the previous 30 days (Table 3). It was then suggested that the chemical nature of lichen surfaces may change with rainfall events, giving rise to a higher particle retention efficiency (14). Other authors have reported losses of elements caused by rainfall in places with long dry deposition periods (80–82). However, these authors did not discriminate between soluble and particulate elements. In general, after a long dry period the first moments of rain produce very concentrated metal solutions, whereas later rainfall produces more dilute solutions. Elements with low affinities for wall binding sites may become bound in the first moments of a rainfall event before being washed. Elements in the precipitation will compete with those already bound to the wall exchange sites. Concentrations of the elements in the rainfall and affinity for the binding sites will be the major factors determining the new chemical composition of the lichen wall after rainfall. In general, it may be concluded that in rainfall events the retention of elements by

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TABLE 3 Correlation Coefficients Between the Concentration of Cu and Fe in Ramalina fastigiata and Usnea spp. During Each Transplantation Period, with Variables Related to Daily Rainfall During the Same Transplantation Periodsa Ramalina fastigiata Rainfall variables (mm) Total rainfallb Average daily rainfall Cumulative rainfall 10 days before transplant retrieval Cumulative rainfall 20 days before transplant retrieval Cumulative rainfall 30 days before transplant retrieval

Usnea spp.

Cu

Fe

Cu

Fe

0.83** 0.83** 0.85**

0.82** 0.82** 0.82**

0.63* 0.64* ns

0.71* 0.72** 0.67*

0.91**

0.88**

0.69*

0.73**

0.95**

0.92**

0.80**

0.83**

** P ⬍ 0.05; * 0.1 ⬎ P ⬎ 0.05; ns ⫽ P ⬎ 0.1; for Usnea spp. n ⫽ 8; for R. fastigiata n ⫽ 7; a The lichen samples were transplanted to ⬃760 m from the center of the mine. b Cumulative rainfall calculated for each transplantation period.

the lichen surface depends on element solubility (depending also on the emission source); particle size; nature of the elements (particularly their binding affinity for the lichen exchange sites); form of the lichen surface (structure and chemistry); and climate (duration of the dry period and the type of rainfall). 5.2

Dry Deposition

Atmospheric dry deposition of metals may occur through sedimentation and impaction (79). In a study of a Cu mine in the south of Portugal with a very dry climate, transplanted lichens intercepted more particles through impaction than by sedimentation (14). Nevertheless, sedimentation probably makes a greater contribution in countries with more humid climates. Copper particles trapped in the deposit dust gauges showed no correlation with Cu intercepted by the transplanted lichens (14,83). The particles measured by deposit dust gauges normally fall into the devices as a result of their own settling speed or are carried in with precipitation. Furthermore, these deposit dust gauges are only satisfactory for the collection of particles greater than 100 µm diameter; smaller particles usually only settle by impacting (15). This kind of gauge tends to underestimate deposition during high wind speed periods (15). However, concentrations of the same element intercepted by directional dust gauges significantly correlated with the total Cu intercepted by the lichens (14,83). The directional gauges were capable

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of collecting not only direct settling dust but also wind-driven particles. Thus, in a dry climate lichens intercept most of the elements by impaction. Several authors state that dry deposition is substantially greater at the end of the summer, having accumulated during the dry period (14,80,84). 5.3

Substrate Sources

Although lichens have long been used as biomonitors, studies of element sources other than air pollution have been limited. Recent research on uptake by lichens suggested that lichens may take up elements from the substrate (85–87). Lichens may cause weathering of rock surfaces by both mechanical and chemical processes. After nutrients have been solubilized, elements may be taken up. Lichens growing on certain metal-rich rocks produce secondary substances—lichen acids—to complex metal ions. Lichens produce a wide range of secondary products, many of which occur as extracellular crystals within the lichens. Most are exclusive products of lichen symbiosis; some are useful in medicine and cosmetics. Purvis et al. (88) showed that lichens growing in Cu-rich substrates (5% Cu dry weight) produced norstictic acid and psoromic acid within the lichen thalli to avoid the toxic effects of Cu. Lichens were found to grow directly on uranium minerals and concentrate uranium within their tissues by McLean et al. (89), who suggested that melanin-like pigments were involved. Other authors have found significant differences in the concentrations of some elements between Flavoparmelia baltimorensis growing on granite and tree bark (86). For a review on lichens growing in metal-enriched environments, see Purvis and Halls (90). Contamination from the soil underneath terricolous lichens can occur by capillary diffusion from more basal tissues, soil or litter (91), or from local windblown particles. In terricolous lichens the bases are more influenced by the soil whereas the state of the tips mainly reflects atmospheric deposition. Epiphytic lichens are subjected only to wind-blown dust; the relative influence of these particles depends on the height at which they grow. On a local scale, particles derived from factories, mines, and traffic are important for monitoring heavy metal deposition. At background areas the particles measured in biomonitors are simply soil or rock dust carried out by wind, rain splash, or animal activity (92). Goyal and Seaward (85) showed that rhizinae of Peltigera canina are capable of absorbing, accumulating, translocating, and regulating soluble metals, which then are able to move freely to and from the upper thalli surface. The poikilohydrous water relations of lichens, with cyclic wetting and drying, presumably accelerate the uptake of soluble mineral material from the substrate by the diffusive mechanism (91). Iron, aluminum, and titanium are often used as indicators of soil contamination (16,87,93,94). Under natural conditions, elevated levels of these elements

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appear mainly to indicate local wind-blown dust, particularly in areas with surface soil disturbance (93,94). In very dry and unvegetated regions, a stronger dispersal of these elements occurs, and direct comparison of metal concentrations and biomonitor results between different climates is not possible. In fact, Loppi et al. (87) showed that in remote areas of the Mediterranean the elemental composition of unwashed epiphytic lichen samples was affected by soil contamination; lithogenic elements (Al, Fe, and Ti) in soils were correlated with the same elements in the lichens. They also concluded that dust contamination is highly variable and probably depends on local site characteristics. Elements dissolved in precipitation may be taken up by epiphytic lichens through scavenging from the atmosphere or leaching of the canopy or tree trunk where the lichen stands. Moreover, the growth form or the position in the tree of a lichen may influence its interception of elements. A fruticose lichen is in less contact with the substrate than, for example, a foliose or a crustose. On the other hand, lichens located in the runoff area of a tree trunk and lichens located at the ends of small branches are exposed to different levels of elements. Farmer et al. (95) showed that epiphyte tissue chemistry responded to seasonal changes in stem flow chemistry. Different tree canopies have different abilities to intercept metals, so that lichens growing in different tree species are submitted to different chemical environments. Some elements (N, K, Ca, Cu, Fe, and Zn) in lichens growing in several Quercus species were shown to be influenced by the substrate element content (86). Oliveira et al. (51) found that epiphytic lichens collected from different phorophytes (tree bark) showed significant differences in their contents of Ca, Mg, K, Zn, and Fe, though no significant differences between K concentrations were observed. Potassium is mainly located within the cell and thus less related to environmental conditions. The differences observed for the other elements could be related to differences between canopy architectures. Variations in bark pH might also be responsible for some of the observed differences between the element contents of lichen samples (95,96). 5.4

Influence of Lichen Interception Surfaces

Branquinho (14) reported that the capacity of Ramalina fastigiata to intercept Cu was approximately twice that of Usnea spp., due to a greater efficiency in intercepting particles. The proportion of soluble to particulate Cu depended on the lichen species and was a reflection of the capacity of each species to differentially take up soluble or particulate Cu. There are few studies on the efficiency of capture of various particles by different plant surfaces (15). It is frequently claimed that differences between collecting surface characteristics, such as roughness, explain differences in the capture and retention of metal-rich particles; hence deposition onto pubescent leaves is expected to be greater than deposition onto smooth leaves (15). Fine hairs on the surfaces of leaves project through the

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viscous sublayer and trap particles that would otherwise have insufficient inertia to reach the vegetation surface (15,96,97). The results of the study of Branquinho (14) were unexpected, since Usnea spp. shows greater hairiness than R. fastigiata. Higher efficiency of collecting particles is also normally explained by higher surface volume ratios of thalli (15,91), which appeared to be greater in Usnea spp. than in R. fastigiata. Thus, in this case, it was suggested that the ability of the different lichen species to intercept particles might be related to chemical, adhesive, and water retention characteristics of the thalli rather than structural features (15,97,98). Future studies of thallus surface chemistry should elucidate differences in lichens’ efficiencies of interception and retention of particles from dry deposition. 5.5

Metal Release Processes

Lichens may release metals by losing biomass through fragmentation by wind, consumption by arthropods, or release of reproduction structures (isidia and soredia). Unbound elements on the lichen surface may be lost during rainfall periods (80,81). Redistribution of elements bound to the wall occurs through input of elements from rainfall dry deposition and particle dissolution. Particles may give rise to soluble material, though to date no studies have determined rates of release of metals from lichens. De´ruelle (99) showed that Pb acquired by lichen transplants at sites contaminated with automobile exhaust was lost within months once they had been returned to their original uncontaminated site. 6

UPTAKE AND CELLULAR LOCATION OF ELEMENTS IN LICHENS

The analysis of air pollution uptake by biomonitors is important because only those pollutants that are absorbed in living cells will directly affect metabolism (100). In addition, analysis of air pollution uptake can provide insight into the air pollution sink strengths of ecosystem components. 6.1

Metal Uptake

The mechanisms of metal uptake in lichens have been reviewed by several authors (9,14,38,91,101–103) and can be summarized as follows: 1. Extracellular ion exchange. Early studies on metal uptake by lichens emphasize the fact that it mainly occurs at extracellular sites through cation exchange processes. These extracellular sites are presumed to be in the cell walls and on the outer surface of the cell membrane (91,102), and it has been suggested that the lichen thallus is similar to an ion exchange resin (29,104,105). The lichen wall binding sites are

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probably carboxylic acid in nature, and part of the protein component of the fungal cell walls and membranes (106). This process is physicochemically regulated, rapid, and reversible. Dissolved metal ions are rapidly bound to extracellular sites (38). 2. Intracellular uptake. This requires the passage of an element across the plasma membrane, using an appropriate carrier system. Uptake is slower and more selective than by ion exchange, and the rate of uptake remains approximately linear for a longer period than for extracellular uptake. 3. Particle entrapment. Trapping of particles contributes significantly to the elemental levels found in lichens (91). Particles are accumulated onto and within the lichen, and may later be solubilized to some extent by secondary lichen products. The occurrence of metals in particles in lichen thalli has been demonstrated by Garty et al. (28). Partial dissolution, binding to exchange sites, and incorporation into the cell are all possible (43,45). It has been suggested that the accumulation and retention of high concentrations of metal pollutants by lichens can be explained by their high cation exchange capacity and morphological features (surface, structure, and roughness), which allow interception and retention of particles (11,16,103). Most studies dealing with lichens as biomonitors have up to now provided evidence about the total amount of metals accumulated in the lichen thallus. The elements in the particulate form will have a short-term impact only if some particulate dissolution occurs, giving rise to soluble material. On the other hand, soluble elements, if located extracellularly, are assumed to be less important for immediate metabolic effect than those located intracellularly (5,102). Therefore, the potential of any element to affect metabolic processes directly cannot be assessed by an analysis of whole-lichen element concentrations (44,74,107). Hence, the fact that a thallus has a high content of a toxic element does not necessarily indicate that the organism possesses some cellular tolerance mechanism. Understanding the availability to the lichen (estimated through quantification of element solubility) and the cellular location of potentially toxic compounds is therefore an important step for establishing whether or not such elements affect cellular mechanisms (102). It is also essential that lichen responses to a particular pollutant or pollutant mixture are clearly defined and understood before setting out a field experiment (18). To this end, considerable controlled plant screening to establish dose-response relationships is required. 6.2

Sequential Elution Techniques

Quantification of metals associated with extracellular exchange sites on the cell wall of plants, fungi, bryophytes, and lichens has been achieved through cation

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exchange techniques (5,12,54,108). In lichens, the absence of a cuticle allows the location and quantification of extracellular and intracellular metals by a sequential elution technique (12,13,44,46,109). The cations bound to the wall can be displaced efficiently by other cations with a strong affinity for those binding sites (110,111) or by chelating agents (44). The quantities of metals subsequently released correspond to the amounts of intracellular elements retained by the cell membrane (110). A suitable elution method requires that the displacing agent should not cause membrane damage in order to avoid ambiguity as to the location of the added metal (110). Other workers have employed electron microscope techniques, which may involve some redistribution of elements during sample preparation to visualize, but not exactly quantify, elements at the same location. Nickel has been found to be a suitable displacing agent for most ‘‘class A’’ physiologically important elements (e.g., Na, K, Mg, and Ca) and many more ‘‘borderline’’ heavy metals (e.g. Cd, Zn, and Co) (4). Strontium has also been used for the displacement of most of the ‘‘class A’’ elements. Elements with considerably greater ‘‘class B’’ characteristics (e.g., Pb, Cu) can be displaced with Na2-EDTA (dissodium ethylenediaminetetraacetic) (13,14,44). 6.3

Cellular Location of Elements Under Field Conditions

Under field conditions, metals are delivered to the lichens dissolved in precipitation and in particulate form (16). Until now, only a few in practical biomonitoring studies using lichens have discriminated between soluble and insoluble metals, as well as intra- and extracellular location of the soluble metals (14,37). Sequential elution of samples of field-polluted lichens growing near a busy road revealed that most of the metals intercepted by the lichens were located extracellularly in a soluble form (14). Moreover, the total analysis did not provide information on species differences for metal uptake. Determination of the location (extra-/intracellular) and availability (soluble/insoluble) of the elements Pb, Zn, and Cu in the studied lichens allowed detection of differences between species, which were not revealed by analysis of whole-lichen element concentrations. In temporal studies the effects of rainfall events on extracellular element concentrations were shown to be important, since K and Mg appeared to be leached from the lichen thallus, whereas other elements (i.e., Pb, Zn, Cu, and Ca) appeared to remain bound to the lichen (14). Boonpragob and co-workers (80,81) also reported that lichen total metal content shows no seasonal pattern, whereas data on leachable elements showed seasonal variations with pollution. Knowledge of the proportion of physiological elements in the different fractions is necessary to quantify and evaluate possible physiological alterations that might change the response of lichens to their environment and thus affect their biomonitoring capacity. The concentrations of K, Mg, and Ca in either extra- or intracellular locations near a busy road and a Cu mine were measured (14). How-

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ever, only extracellular soluble and particulate K, Mg, and Ca were found to decrease with increasing distance from the pollution sources (14). The soluble intracellular elements remained relatively constant (unless alterations in membrane permeability occurred due to toxic effects) and did not reflect the variations with distance from the source. Thus, even in the case of macronutrients (i.e., K, Mg, and Ca), which have high baseline concentrations, it is possible to evaluate the influence of the environment whenever their location (intra- and extracellular) and form (soluble/particulate) is taken into account (14). The authors of a study of the impact of deposition of saline elements on the southwestern coast of Portugal (84) found that only the concentrations of extracellular ions decreased with increasing distance from the coast. The intracellular fractions of these elements are relatively independent of the ions on the surface and bound to the cell wall due to physiological control by the organisms. The intracellular location of Cu, K, and Mg can explain the physiological changes and species survival in the surroundings of a copper mine (14). The proportions of particulate and extracellular soluble Cu in lichens allowed a comparison with the availability of Cu in the suspected source; discrimination between soluble and particulate Cu and the location of soluble elements at extraor intracellular sites in the lichens provided more relevant information on the nature of the intercepted pollutants and their effects on lichen survival than total analysis. Extracellular and particulate metals are dependent on environmental sources, whereas those located within the cell are more related to physiological processes (107). 7

MECHANISMS OF TOLERANCE TO METAL POLLUTION

Though several hundred publications have dealt with lichens and metals, only a few describe lichen species as pollutant-resistant or pollutant-tolerant, illustrating our lack of knowledge of these organisms’ mechanisms of tolerance to high metal content (112,113). Sarret et al. (113) described the mechanisms of lichen resistance to metal pollution by comparing Diploschistes muscorum, a hyperaccumulator of Pb and Zn, and Xanthoria parietina, a Pb-tolerant species. Under normal conditions D. muscorum produces oxalate, but synthesis is enhanced by exposure to metal pollution, accounting for the hyperaccumulation of Zn and Pb (1.2- to 2-fold) from the substrate (113). It is also suggested that other cations, such as Fe, Cu, Cd, and Ca, may also be complexed to oxalate. Moreover, the authors propose that lichens excrete the oxalates and so this is an extracellular immobilization mechanism. In contrast to D. muscorum, X. parietina does not increase oxalate production in polluted conditions. Sarret et al. (113) concluded that Pb in X. parietina was predominantly complexed to the fungal cell walls by a passive ion exchange process.

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Purvis et al. (88,114) also showed that lichens growing in Cu-rich substrates (with 5% Cu dry weight) produced norstictic and psoromic acids within the lichen thalli for the formation of metal-lichen acid complexes. The presence of lichen acids explains how some lichens may tolerate metals by means of a detoxification mechanism, arising from the formation of an insoluble metal-lichen substance complex. It is generally accepted that X. parietina is able to tolerate high levels of pollutants, including heavy metals (48,67,112,113), and that Usnea spp. is more sensitive to general air pollution (Figs. 4, 8, and 9) (14,37). In order to evaluate the cation exchange capacity of different lichen species, samples of several species were submitted to saturated concentrations of Pb and Cu (Fig. 10), elements with the greatest affinity for the lichens’ wall binding sites (4). The samples were then submitted to a sequential elution to displace Pb and Cu from the extracellular sites, which is a measure of their cation exchange capacity (Fig. 10). The results show that the cation exchange capacity values obtained for the different species were in agreement with the previously determined susceptibility to pollution. Thus, X. parietina had the lower cation exchange capacity, meaning that its toler-

FIG. 10 Lichen cation exchange capacity measured by the Pb and Cu displaced from the extracellular sites after lichen samples of several species have been submitted to Pb and Cu saturation concentrations for 2 h. The extracellular fraction was obtained by elution of the lichen samples with 20 mM dissodium ethylenediaminetetraacetic (Na2-EDTA). Columns are means of 5–10 replicates.

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ance is probably based on a metal avoidance mechanism. On the other hand, Usnea spp. showed the highest cation exchange capacity (Fig. 10), which is also in agreement with its relative susceptibility to metal pollution (13,14,37). The wall-bound metals may enter the cell more easily, causing negative physiological effects. In another experiment, Silberstein et al. (112) compared X. parietina, a resistant lichen, with the more sensitive Ramalina duriaei in relation to possible air pollution protection mechanisms. They found that the former has a multitude of possible protective systems that can be classified in two categories: (a) induced or stimulated by exposure to pollutants; and (b) constitutive defense. The first may occur through (a) oxidation of SO2 to nontoxic sulfate; (b) an increased glutathione content; (c) increase in amino acid synthesis, mainly proline and arginine; or (d) an increase in enzymatic detoxification of active oxygen forms. The second are enabled by (a) efficient buffering capacity, (b) high K content, and (c) antioxidation by parietin. Since X. parietina is a foliose species and R. duriaei is fruticose, the former has a greater ratio of surface exposed to the atmospheric pollution to volume. Lichens subjected to air and soil pollution may have three mechanisms of metal immobilization: (a) entrapment of metal-rich particles, which depends on their surface/volume ratio; (b) intracellular uptake with likely complexation to S-containing peptides and metallothioneins; and (c) extracellular complexation within fungal cell walls, the metal being complexed to hydroxyl, carboxyl, phosphate, amine, or sulfhydryl groups. In the case, the low cation exchange capacity might be a strategy to avoid metals entering the cells. The authors also suggest that metals may be immobilized in lichens by organic acids and lichen substances, particularly the oxalates due to their strong metal binding ability, as is the case of D. muscorum and lichens growing in Cu-rich substrates. For instance, parietinic acid, which is an anthraquinone produced exclusively by X. parietina, possess several functional groups that can bind metals (113). As in higher plants, these ligands may be involved in excretion and immobilization of metals within the vacuole. Apparently, lichens have differing strategies to avoid metals. The mycobiont is the most important partner in terms of metal accumulation, and it seems that more comparisons with other biological models such as mycorrhiza could benefit lichen biomonitoring techniques. 8

SOURCES OF VARIABILITY WHEN SAMPLING LICHENS

Another drawback to the use of biological systems as monitors is that their reaction depends on their specific characteristics (genetic, age, state of health, type of reproduction, etc.) and on environmental parameters (elemental availability, substrate, topography, micro-/macroclimatic conditions, etc.) (7,18). For this rea-

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son the standardization of biological monitors is difficult (7). Lichens are one of the most frequently used organisms for biomonitoring of atmospheric pollution. However, there is a lack of knowledge of the causes of sample variability, which is an important source of error in biomonitoring analysis (6,20). Before starting a biomonitoring survey, it is necessary to test the effect of different phorophytes on levels of elements in the epiphytic lichens growing below (51,86). This should cover canopy effects and the relative position of the lichen in the phorophyte (see Sec. 5.3). If differences are observed between phorophytes, care should be taken to choose only one or to intercalibrate lichens concentrations between phorophytes. Atmospheric heavy metal pollution gives rise to anomalous high concentrations of metals in the environment, originating from a variety of anthropogenic sources as well as natural geochemical processes. Heavy metals occur naturally in rock-forming and ore minerals, and so a range of normal background concentrations of these elements can be found in soils, sediments, water, and living organisms. Lichens may also be contaminated by wind-blown dust from the surrounding soils (see Sec. 5.3). However, these background values may very substantially at local, regional, or national scales. Soil in areas rich in minerals (e.g., serpentine and calcareous areas) may contain high levels of metals. Lichens may also become naturally polluted in volcanic areas or on coastal regions. Grasso et al. (115) found that levels of As, Sb, Br, and Pb in lichens collected in the volcanic areas of Mt. Etna and Vulcano Island in Italy reflected the local gaseous emissions (plume and fumaroles). Thus, for all studies, the background levels of pollutants must be evaluated at the study site; comparison with other sites in the region may not be appropriate. It is important to consider that sampling of lichens for biomonitoring studies may take a few days, if not weeks, due to the high number of sampling points normally involved. It is advisable to perform the sampling during the shortest time possible under stable climatic conditions. Rainfall or a long dry deposition period (see Secs. 5.1 and 5.2) might substantially change the concentrations of elements in lichens (14,51,80,81,84,116), such that spatial comparisons are not possible. When the pollutants under consideration also have an important physiological role, it may be important to discriminate between insoluble, soluble extracellular, and soluble intracellular compounds in order to distinguish between environmental sources and physiological requirements (see Sec. 6). Several authors have shown that metal deposition is usually a local phenomenon (37,67,84), whereas gaseous deposition occurs over longer ranges. Several sampling networks have distances between sampling points of 10–30 km (117,118). This type of sampling network does not allow identification of all the possible sources of metallic pollution. Sites where the local deposition of metals may occur include roads (99), mines (14,37), and cement mills (67). The possibility of unknown sources of pollution being identified by such a network is very

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small, considering that most of the metal pollution sources previously described have an impact radius not greater than 2 km. Thus, a sampling network to measure metallic atmospheric deposition should be based on local factors, such as location of known sources, topography, altitude, predominant winds, etc., and should be designed so as to allow subsequent identification of the known and unknown pollution sources. Some of the problems described above can be ameliorated by the use of transplanted lichens, which are homogeneous because they are all collected from the same location. They are especially useful for biomonitoring studies where lichens are not normally present, or when the study of the dynamics of deposition is intended (16). Sampling of lichen transplants is useful in pinpointing sources of pollution and can reveal dispersion patterns of particulate material (15,57,99). When sampling lichens for biomonitoring purposes it is important to consider the age of the thalli to be analyzed. Since there is no way of identifying the exact age of a certain part of a lichen, we suggest testing whether there are differences between young and old parts of the lichen thalli before implementing biomonitoring programs. Oliveira et al. (51) established the relative importance of several sources of variability in sampling lichens for biomonitoring: variations in the chemical composition of the lichen thallus with ages; and the influence of the substratum in chemical composition of epiphytic lichens (growing on tree bark). Levels of Ca, Mg, and K in marginal and central zones of the lichen thallus exposed to different pollution conditions were compared, taking into account the solubility and the location of the elements in lichen thalli (Fig. 3). The central zones of the thallus had significantly higher concentrations of particulate Ca, Mg, and K than the marginal ones (Fig. 3). Central parts, which are older, had been exposed to dust particles for a longer period. This could be related to the continual entrapment of particles by fungal hyphae during lichen growth. Lichens collected near a cement mill contained higher concentrations of extracellular soluble elements than those from an urban environment (Lisbon) (Fig. 3); dust production may have caused a higher deposition of elements on the lichen surface at the cement mill. It has been reported that some elements may be released from inside the lichen cell, particularly K due to desiccation damage (41). However, intracellular K concentrations in lichens from the cement mill were similar to those from Lisbon, suggesting that membrane damage had not occurred (Fig. 3E, F). Variations in intracellular concentrations are mainly related to physiological requirements and are more difficult to interpret in terms of environmental sources (14) (see Sec. 6.2). 9

CALIBRATION LICHENS WITH DUST GAUGES

Another difficulty in the use of lichens as biomonitors is that it has very rarely been possible to find a linear relationship between concentrations of pollutants in lichen thalli and in the atmosphere (119). This is important because regulations

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always define threshold levels in terms of atmospheric concentrations. Moreover, the application of biomonitors has several advantages over the use of physical measurements: common occurrence in the field; easy sampling and transplanting; greater aptitude for pollutant accumulation; and no requirement for expensive technical equipment (7,15). However, in order to determine absolute atmospheric concentrations, the relation between the concentration of elements intercepted by the lichens and atmospheric deposition rates must be calibrated. Some qualitative (93,116,120,121) and quantitative (14,82,83,122–126) comparisons of biomonitoring methods with physical and chemical measurements of atmospheric deposition have been published.

FIG. 11 Plots of observed vs. predicted values obtained by linear multiple regression models between Cu concentration in R ⫽ R. fastigiata (A and C) and U ⫽ Usnea spp. (B and D) and total volume of rainfall, TR (C and D) with the dependent variable, Cu concentration intercepted by the directional gauges (Dir): A (Dir ⫽ 0.83 ⫹ 17.40*R); B (Dir ⫽ 7.52 ⫹ 28.25*U); C (Dir ⫽ ⫺3.14 ⫹ 14.43*R ⫹ 0.13*TR); and D (Dir ⫽ 3.81 ⫹ 24.67*U ⫹ 0.11*TR). (From Ref. 83.)

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In order to transform the concentrations of Cu and Fe in lichens (expressed in µmol g⫺1 dry weight) into atmospheric deposition values (mg m⫺2), a calibration between the Cu concentrations in lichens and dust gauges was attempted for comparable periods (14,83). Levels of Cu in the directional gauges were significantly correlated with the total Cu intercepted by the lichens (14,83), a significant fraction of which was of particulate origin. The introduction of rainfall variables improved the regression model, demonstrating that element levels in lichens were dependent on this environmental factor. Among the rainfall-related variables, total rainfall and average rainfall were the ones with the greatest influence in Cu deposition (see Sec. 5.1) (14,83). A calibration model was obtained that permits the transformation of element concentrations in lichens (µmol g⫺1) into atmospheric deposition rates (µmol m⫺2 month⫺1). The model performance measured by observed vs. predicted values seems quite good, considering that biological material was involved and that the variability of data from biological systems is normally very high (Fig. 11). 10 FINAL REMARKS As seen before, lichens are extremely complex organisms and their basic biology is still not totally clear. However, when using them as biomonitors there is a tendency to simplify such complex systems. I would like to point out that reliable data based on biomonitors must be supported by more research, particularly in the following areas: 1. It is suggested that researchers do more work on the location of metals in lichens to allow determination of the threshold intracellular concentrations above which metals cause physiological changes in lichens. On the other hand, both the insoluble (particulate material) and the extracellular (surface and wall bound) metal concentrations are more dependent on environmental exposure than on total concentrations in the lichens. 2. Careful planning is required to avoid the high variability associated with sampling lichens for biomonitoring purposes. However, the degree of error associated with biological variability can be offset by the use of a large number of sampling points, permitting the production of reliable pollution models that could not be obtained by instrumental recording owing to high unit costs and the scarcity of the sampling points. 3. Use of lichen biodiversity data to map the impact of atmospheric pollution should be accompanied by the concentrations of elements accumulated in the lichen thallus to allow distinction between climatic and pollution factors. Finally, models of the interaction of metals with lichens should be confirmed under field conditions; laboratory-based data may be unrealistic.

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4. Lichens are able to integrate their physical, chemical, and biological exposure to atmospheric metals. The fact that biomonitors integrate this information should be considered an advantage because it takes into account the interaction between different kinds of pollution. The use of biological monitors is more representative of the field conditions and of the ecosystem than physical monitoring devices such as deposit gauges (15). Dust gauges or high-volume samplers emphasize the physical aspects of the pollution and do not consider the chemical and biological aspects, which in some cases are more significant (19). Nevertheless, whenever possible the concentrations of elements in lichens should be translated into atmospheric deposition rates to allow comparison with legislation based on dust gauges or high-volume samplers. ACKNOWLEDGMENTS I would like to acknowledge Dr. D. H. Brown and Prof. F. Catarino for discussing some of the data presented here. I would like to acknowledge the helpful comments made on a draft of this manuscript by Drs. Cristina Ma´guas, Gisela Oliveira, and Steve Houghton. REFERENCES 1. TH Nash III. Lichen biology. Cambridge, UK: Cambridge University Press, 1996, pp. 1–303. 2. V Ahmadjian. Lichens are more important than you think. Bioscience, 45:124, 1995. 3. R Honegger. Morphogenesis. In: TH Nash III, ed. Lichen Biology. Cambridge, UK: Cambridge University Press, 1996, pp. 24–36. 4. E Nieboer, HS Richardson. The replacement of the nondescript term ‘‘heavy metals’’ by a biologically and chemically significant classification of metal ions. Environ Pollut (Series B) 1:3–26, 1980. 5. GM Gadd. Interaction of fungi with toxic metals. New Phytol 124:25–60, 1993. 6. B Markert. The biological system of the elements (BSE) for terrestrial plants (glycophytes). Sci Tot Environ 155:221–228, 1994. 7. R Wittig. General aspects of biomonitoring heavy metals by plants. In: B Markert, ed. Plants as Biomonitors: Indicators for Heavy Metals in the Terrestrial Environment. New York: VCH, 1993, pp. 3–28. 8. HW Woolhouse. Toxicity and tolerance in the response of plants to metals. In: OL Lange, PS Nobel, CB Osmond, H Ziegler, eds. Responses to the chemical and biological environment. Berlin: Spring-Verlag, 1983, pp. 245–300. 9. M Galun, R Ronen. Interaction of lichens with pollutants. In: M Galun, ed. Handbook of Lichenology. Boca Raton: CRC Press, 1988, 3:55–72. 10. JP Bennett, MJ Buchen. Bioleff: three databases on air pollution effects on vegetation. Environ Pollut 88:261–265, 1995.

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97. KJ Puckett, EJ Finegan. An analysis of the element content of lichens from the Northwest Territories, Canada. Can J Bot 58:2073–2089, 1980. 98. LJR Boileau, PJ Beckett, P Lavoie, DHS Richardson, E Nieboer. Lichens and mosses as monitors of industrial activity associated with uranium mining in northern Ontario, Canada. Part 1: Field procedures, chemical analysis and interspecies comparisons. Environ Pollut (Series B) 4:69–84, 1982. 99. S De´ruelle. L’utilisation des lichens pour la detection et l’estimation de la pollution par le plomb. Bull d’Ecologie 15:1–6, 1984. 100. WE Winner. Mechanistic analysis of plant responses to air pollution. Ecol Appl 4:651–661, 1994. 101. Y Tuominen, T Jaakkola. Absorption and accumulation of mineral elements and radioactive nuclides. In: V Ahmadjian, V Hale, eds. The Lichens. New York: Academic Press, 1973, pp. 185–223. 102. DH Brown. Mineral uptake by lichens. In: DH Brown, DL Hawksworth, RH Bailey, eds. Lichenology: Progress and Problems. New York: Academic Press, 1976, pp. 419–439. 103. DH Brown. Lichen mineral studies—currently clarified or confused? Symbiosis 11:207–223, 1991. 104. E Nieboer, P Lavoie, RLP Sasseville, KJ Puckett, HS Richardson. Cation-exchange equilibrium and mass balance in the lichen Umbilicaria muhlenbergii. Can J Bot 54:720–723, 1976. 105. E Nieboer, KJ Puckett, B Grace. The uptake of nickel by a physicochemical process. Can J Bot 54:724–733, 1976. 106. DHS Richardson, S Kiang, V Ahmadjin, E Nieboer. Lead and uranium uptake by lichens. In: DH Brown, ed. Lichen Physiology and Cell Biology. New York: Plenum Press, 1985, pp. 227–246. 107. DH Brown, RM Brown. Mineral cycling and lichens: the physiological basis. Lichenologist 23:293–307, 1991. 108. DH Brown, G Brumelis. A biomonitoring method using the cellular distribution of metals in mosses. Sci Tot Environ 187:153–161, 1996. 109. DH Brown, RP Beckett. Uptake and effect of cations on lichen metabolism. Lichenologist 16:173–188, 1984. 110. DH Brown, JM Wells. Sequential elution technique for determining the cellular location of cations. In: JM Glime, ed. Methods in Bryology. Nichinan: Hattori Botanical Laboratory, 1988, pp. 227–233. 111. JM Wells, DH Brown, RP Beckett. Kinetic analysis of Cd uptake in Cd-tolerant and intolerant populations of the moss Rhytiadiadelphus squarrosus (Hedw.) Warnst and the lichen Peltigera membranacea (Ach.) Nyl. New Phytol 129:477– 486, 1995. 112. L Silberstein, BZ Siegel, SM Siegel, A Mukhtar, M Galun. Comparative studies on Xanthoria parietina, a pollution-resistant lichen, and Ramalina duriaei, a sensitive species. II. Evaluation of possible air pollution-protection mechanisms. Lichenologist 28:367–383, 1996. 113. G Sarret, A Manceau, D Cuny, C Haluwyn, S De´ruelle, JL Hazemann, Y Soldo, L Eybert-Be´rard, JJ Menthonnex. Mechanisms of lichen resistance to metallic pollution. Environ Sci Technol 32:3325–3330, 1998.

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114. OW Purvis, JA Elix, KL Gaul. The occurrence of copper-psoromic acid in lichens from cupriferous substrata. Lichenologist 22:345–354, 1990. 115. MF Grasso, R Clocchiatti, F Carrot, C Deschamps, F Vurro. Lichens as bioindicators in volcanic areas: Mt. Etna and Vulcano Island (Italy). Environ Geol 37:207– 217, 1998. 116. A Rodrigo, A Avila, A Go´mez-Bolea. Trace metal contents in Parmelia caperata (L.) Ach. Compared to bulk deposition, throughfall and leaf-wash fluxes in two holm oak forests in Montseny (NE Spain). Atm Environ 33:359–367, 1999. 117. A Ru¨hling. Atmospheric heavy metal deposition in Europe: estimations based on moss analysis. NORD 1994, 9:1–53. 118. MC Freitas, MA Reis, LC Alves, HTh Wolterbeek. Distribution in Portugal of some pollutants in the lichen Parmelia sulcata. Environ Pollut 106:229–235, 1999. 119. PL Nimis, M Castello, M Perotti. Lichens as bioindicators of heavy metal pollution: a case study at la Spezia (N Italy) In: Markert B, ed. Plants as Biomonitors: Indicators for Heavy Metals in the Terrestrial Environment. New York: VCH, 1993, pp. 265–284. 120. P Little, MH Martin. Biological monitoring of heavy metal pollution. Environ Pollut 6:1–19, 1974. 121. Z Jeran, AR Byrne, F Batic. Transplanted epiphytic lichens as biomonitors of air contamination by natural radionuclides around the Zirovski VRH uranium mine, Slovenia. Lichenologist 27:375–385, 1995. 122. A Anderson, MF Hovmand, IB Johnsen. Atmospheric heavy metal deposition in the Copenhagen area. Environ Pollut 17:133–151, 1978. 123. K Pilegaard. Heavy metals in bulk precipitation and transplanted Hypogmnia physodes and Dicranoweisia cirrata in the vicinity of a Danish steelworks. Water, Air and Soil Pollution 11:77–91, 1979. 124. NK Vestergaard, U Stephansen, L Rasmussen, K Pilegaard. Airborne heavy metal pollution in the environment of a Danish steel plant. Water, Air Soil Poll 27:363– 377, 1986. 125. R Figueira, AMG Pacheco, AJ Sousa, C Branquinho, F Catarino. Assessment of two epiphytic lichens as saltfall biomonitors: calibration of transplants. Ecotoxicol Environ Restor 1:61–69, 1998. 126. JE Sloof. Lichens as quantitative biomonitors for atmospheric trace-elements deposition, using transplants. Atm Environ 29:11–20, 1995.

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6 Bryophytes and Pteridophytes Nicholas W. Lepp Liverpool John Moores University, Liverpool, United Kingdom

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INTRODUCTION

There are several well-known associations that occur between mosses, liverworts, and ferns that grow on metal-enriched substrates. These groups of plants each possess certain anatomical and physiological properties that enable some representatives to occupy unique ecological niches in natural and man-made metaliferous environments. The best documented of these are the groups of specialized bryophytes that are found on substrates enriched with copper; so-called copper mosses are found worldwide (1) and come from widely separated taxonomic groups. Other bryophytes are associated with lead- and zinc-enriched substrates. Serpentine soils do not appear to possess a distinctive bryophyte flora. Some species of pteridophytes are associated with serpentine substrates in various parts of the world (2–4), while several African species grow on copper-enriched soils or on soils polluted by metal smelter emissions (5). The existence of these diverse groups is a clear indication of the ability of Bryophytes and Pteridophytes to adapt to extremes of metal content of their growth substrate, either by avoiding

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the toxic constituents or by expressing resistance/tolerance to metals at an organismal or cellular level. 2

BIOLOGY OF BRYOPHYTES AND PTERIDOPHYTES IN RELATION TO POTENTIAL FOR METAL TOXICITY AND TOLERANCE

Because both groups of plants possess motile gametes, they require water for sexual reproduction. In addition, both possess two distinct growth forms (protonema and gametophyte/sporophyte in bryophytes, prothallus/gametophyte in pteridophytes) and have less developed water relations than vascular plants. Each of these properties makes bryophytes and pteridophytes potentially more vulnerable to metals in their growth substrates than vascular plants. Bryophytes do not acquire inorganic ions and complexes by means of root uptake. No species of foliose bryophyte possess a cuticle; leaves are generally a single cell in thickness and desiccate readily. Plants are attached to the substrate by means of nonabsorptive rhizoids, which serve no other function than attachment. Uptake of water and dissolved solutes occurs over the whole plant surface, with the bulk of inorganic solutes arriving via precipitation enriched with foliar wash-off or leachates from plant canopies above the bryophytes. Bryophytes show two different patterns of water uptake, depending on their growth form. The branched, mat-forming pleurocarpous species are ectohydric, taking up water over the whole body surface. In contrast, the upright, nonbranching, tuft-forming acrocarpous species take up water on the outer surface of the stem, with water moving up or down by capillary action. There are no selective barriers to metal uptake in bryophytes, except for the cell walls. Substrate metal content only plays a role in determining bryophyte distribution where soluble metals have direct access to biological processes in the cell cytoplasm or where there is the direct potential for interference with either the sexual reproductive process or the establishment of asexual propagules (gemmae, tubers) (6). Pteridophytes have different water relations to bryophytes. In addition to possessing well-developed cuticles, they possess root systems that absorb water and dissolved ions from the soil and efficient conducting systems that connect roots and shoots. In these respects they are similar to higher plants. However, pteridophytes still depend on water for sexual reproduction, with the sex organs developing on prothalli which, in turn, have developed from spores shed by the mature plant. A film of water is essential for reproduction to take place, so that soluble metals have the potential to impact pteridophytes at a critical stage in their life cycle, despite the possession of similar barriers to unlimited metal access to those found in higher plants. Asexual reproduction is less common in pteridophytes than in other lower plant phyla (7).

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MOSSES AND LIVERWORTS ASSOCIATED WITH METAL-ENRICHED SUBSTRATES

Several species of bryophytes can be classed as obligate metallophytes, their distribution being inextricably linked to soils or substrates rich in various heavy metals. The best known of these are the so-called copper mosses, a group of acrocarpous mosses and liverworts that appear to be exclusively associated with substrates enriched with copper and other heavy metals. These species are either localized endemics or possess worldwide highly disjunct distribution, both reflecting the uncommon nature of their preferred substrates (8). The most widely investigated copper mosses are those in the genera Mielichhoferia and Scopelophila (Merceya). These are associated almost exclusively with copper-rich or other metal-rich substrates, either natural outcrops or substrates enriched as a result of human activity. Of the two genera, Scopelophila is most frequently and incontrovertably associated with metal-enriched substrates. The six known species are almost certainly all copper mosses and S. ligulata is always reported as associated with metal deposits; these include copper, iron, and silver, as well as soils associated with sulfur-enriched waters from hot springs (9). In Europe, S. cataractae is associated with soils contaminated by metal processing industries, especially those that processed sulfide-based ores. All of the habitats at which this species occurs in Europe are artificial, e.g., spoil heaps, soils contaminated with metal particulates or slags or railway embankments, or ballast associated with metal-processing industries. All European populations also appear to be sterile or exclusively male. The substrates with which S. cataractae is associated are also less acidic (pH 5.6–6.8) than those reported for other copper mosses (10). Other populations of S. cataractae have been studied in Japan. Here the species is associated with copper mines in Honshu and Shikoko as well as with some surprising artificial substrates. Populations are recorded growing on rocks receiving drainage from copper roofs as well as occurring on the base of a bronze statue of Buddha (11). Elsewhere in Japan, a single natural population has been recorded on an outcrop rich in limonite (iron oxide), but other occurrences not directly connected with copper-enriched substrates are in the vicinity of hot springs (11). Regrettably, there are no detailed chemical analyses of the substrates that support S. cataractae in those sites not directly contaminated by human activity. The other European species, S. ligulata, has a more restricted distribution in the Alps and Pyrenees (10). There is debate as to the copper requirements of the other genus of copper moss Mielichhoferia. Both M. elongata and the closely related M. nitida have been widely reported to be associated with copper deposits, especially pyrites and its weathering products, in several parts of Europe. Some analysis of substrates associated with these species indicates the presence of elevated copper concentra-

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tions, but in other cases, concentrations are not remarkable. North American species (M. mieilichhoferi, M. macrocarpa) may occur on substrates with some degree of mineralization, but the 14 species of Mielichhoferia found in South America (mostly in the Andes) do not appear to be associated with copper enrichment. Japanese species have not been reported to be associated with copper-rich substrates, but for both South America and Japan, there is a lack of recent chemical analytical data of both plants and soils. In common with Scopelophila, Mielichhoferia species are also associated with sulfur-enriched waters and substrates. Whether this is a primary determinant of their distribution or a secondary aspect, as many of the heavy metals have been derived from sulfide ores, is not clear (1). Two species of the genus Ditrichum, both endemic to the United Kingdom, are totally confined to metal-enriched substrates. D. cornubicum, found in only two localities in Cornwall (southwestern England) occurs on copper-enriched substrates at disused mine sites, whilst D. plumbicola occurs on lead mine spoil in Northern Wales, Northumberland, and the Isle of Man (12,13). As with European populations of S. cataractae, both of these species have never been found in a fertile state, but D. cornubicum is known to aestivate by means of tubers. The moss Grimmia atrata is also closely associated with acidic, copper-bearing rocks in the United Kingdom and elsewhere in Europe and Japan (12). Several species of the leafy liverwort Cephaloziella are also closely associated with metal-enriched soils and substrates. C. nicholsonii is a UK endemic species, confined to old copper mine workings in southwestern England and Wales, while the closely related C. masalongi is confined to similar substrates but with a wider distribution in Europe and North America (14). UK populations of C. integerrima are confined to copper soils in Cornwall, but the species has a wider European distribution, whereas the more common C. rubella, C. hampeana, and C. stellulifera are all found on old mine spoils in the United Kingdom. Neither C. nicholsonii nor C. masalongi are reported as fertile in their UK sites (14). Both species produce gemmae, so asexual reproduction may have perpetuated apparent copper dependency in both species. The leafy liverwort Gymnocolea acutiloba has also been reported to occur predominantly on copper-enriched substrates in several parts of the world, but its single UK location is nonmetalliferous (13). There is a possibility that this may not be a distinct species. Studies on bryophytes associated with areas of copper mineralization and processing have been reported from Shaba province, Zaire (15). Three species (Brachymenium acuminatum, B. philonotula, and Campylopus bequartii) occurred most frequently, while other taxa (Bryum arachnoideum and unidentified species of Bryum and Pottia) occurred much less frequently. There is no indication if any species in the first group are exclusively cuprophilic. Shacklette (16) described in some detail the bryophyte communities associated with various types of mineralization in Alaska. Two species, the leafy liver-

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wort Nardia scalaris and the acrocarpous moss Oligotrichum hercynicum, were found to occur in pure stands on several areas of copper-enriched soil (0.2–0.6% Cu), while the leafy liverwort Gymnocolea acutiloba was only found to occur on such substrates. An exposure of galena/sphaelerite ore (PbS/ZnS) was found to be exclusively colonized by the leafy liverwort Cephalozia bicuspidata. No significant bryophyte associations were reported from the vicinity of old Mercury (cinnabar) mining activity. Despite the wealth of information on the unique assemblages of vascular plants that occur on serpentine substrates, there appear to be no serpentinespecific bryophytes reported in the literature. Indeed, a depauperate bryoflora has been reported for at least one UK serpentine site at Keen of Hamar, Shetland (17). In many instances, bryophytes associated with metal-enriched substrates are also abundant on other substrates. Several ubiquitous urban taxa (Bryum argenteum, Funaria hygrometrica, and Pohlia nutans) are frequently found on metal-polluted soils. P. nutans is particularly associated with natural and polluted copper-rich soils. It has been reported to form a pure turf in a copper swamp forest in New Brunswick, Canada (18) in situations where total soil copperconcentrations were 3–10%; trees and other vascular plants found these conditions phytotoxic. This species was also a dominant constituent of the bryoflora in copper-contaminated turf at the BICC copper rod plant at Prescot, Merseyside, United Kingdom (19), occurring where tolerant grasses (Agrostis sp.) failed to flourish. Bryum argenteum has also been reported to flourish on copper-rich substrates (20). There have been few systematic field investigations of the direct effects of heavy metals on the ecology of bryophytes. In a field study at a well-characterized site (the BICC copper rod plant, Prescot), Lepp and Salmon (19) found clear differences in the bryoflora that could be attributed to differences in soil copper concentrations. Where these exceeded a threshold (total HNO3-extractable Cu ⬎ 350 mg kg⫺1 dry wt), pleurocarpous moss species disappeared, to be replaced by acrocarpous species (Pohlia nutans, Barbula recurvirostra, and Bryum rubens—the latter where the substrate was alkaline due to the presence of limestone chippings). This was attributed to either a reduction in grass cover, leaving gaps in a closed sward for invasion by more metal- and drought-resistant mosses, or reduced potential for metal access due to the ectohydric water relations of the acrocarpous species. In addition, routine mowing of the grassland would have prevented colonization of bare patches by annual or perennial phanaerogams. It would appear that where soil metal concentrations have a directly adverse effect on higher plants, some bryophyte species are clearly able to colonize and thrive on contaminated substrates. These do not tend to be species confined to such substrates, but a number of common mosses and leafy liverworts appear to be competitive in such situations.

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(a)

(b) FIG. 1 (a) Bioconcentration of different elements in Scapania undulata (mg/ kg dry wt) before and after bioassay with when cultivated in solutions containing 70–100% sewage accumulated lead 85 times in 100% sewage and 58 times in 70% sewage. Mercury content increased 40 times in 100% sewage and 20 times in 70% sewage. (b) Folds of bioconcentration of different metals. (From Ref. 21.)

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Experiments with the liverwort Scapania undulata collected from a clean, montane forest stream when cultivated in solutions containing 70–100% sewage accumulated lead 85 times in 100% sewage and 58 times in 70% sewage. Likewise, mercury content increased 40 times in 100% sewage and 20 times in 70% sewage. Increased contents of cadmium, chromium, copper, and nickel were observed (Fig. 1) (21). It was observed that this liverwort when exposed to 70% sewage did not show significant toxic symptoms and hence suggested its possible use in removal of toxic metals from water. 4

PTERIDOPHYTES ASSOCIATED WITH METAL-RICH SUBSTRATES

There is a dearth of information on this topic. Few systematic investigations have considered the interactions between members of the various groups of pteridophytes and heavy metals, and there are no groups that are comparable with the copper mosses referred to above. Brooks and Malaisse (5) list several species of pteridophyte associated with different types of metal-enriched substrates in southcentral Africa. In Zimbabwe, ferns are very rare on nonserpentine Ni-rich soils; only Pellea calomelanos and Cheilanthes hirta were reported. No ferns were reported from chromium- or arsenic-rich soils, but Pteris vittata may often occur on arsenical mine dumps (22–25). This latter species is also found on arsenical mine dumps in West Africa (26), while As accumulation has been demonstrated in Ceratopteris cornuta from the same area (27). Ferns are more frequently encountered on copper-enriched substrates. Wild (28) reports 10 taxa that occur on such substrates in Zimbabwe, and Jacobsen (29) lists an additional 7 species. These totals make the pteridophytes the fourth most abundant group (in terms of species richness) of plants associated with copper soils. This may be influenced by the rocky nature of many copper-rich sites, allowing ferns to flourish in the shaded and moist crevices that predominate in this type of terrain. A similar argument has been advanced to account for the presence of P. vittata at Ghanaian mine sites, as this species is absent from the remainder of the West African tropical region (26). Studies on the flora of metal-enriched soils in Shaba province, Zaire have demonstrated the significant proportion of the native pteridophytes that grow in such conditions. Of the 90 pteridophyte taxa known to occur in the region, at least 22 occur over mineralized ground, representing 8.4% of the metallicolous flora of the province (5). There is some debate as to the presence of any obligate metallophytes among this group; Cheilanthes inaequalis var. lanopetiolata reportedly confined to copper/cobalt deposits is now considered synonymous with the more widespread and substrate-indifferent var. inaequalis. Mohria lepigera appears to be a locally good indicator of copper deposits in this region, but occurs off copper elsewhere in Zaire (5). Pteris vittata is also abundant on copper-rich

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damp sandy clay soils, frequently forming a monoculture (5). The majority of the pteridophytes associated with metalliferous soils in Shaba province do not appear to accumulate metals in shoot tissues [see table in Brooks and Malaisse (5) for details], with the exception of Ophioglossum lancifolium which appears to accumulate both Co and Cu from its substrate. Ferns are also recorded from serpentine and other ultramafic soils. Wild (28) records seven taxa from the Great Dyke region of Zimbabwe, including Pellea calomelanos and Cheilanthes hirta. However, it would appear that ferns of the genus Asplenium are the most typical group with close associations with such soils in more temperate regions. Two species, the very rare and disjunctly distributed A. presolanense (found in isolated populations in Switzerland and Canada) and the more widespread A. cuneifolium, are considered to be serpentine endemics (4,30). While A. cuneifolium is absent from Britain, there is a very characteristic assemblage of other Asplenium spp. that are found at serpentine and other metal-enriched sites. The most common pteridophyte at UK serpentine sites is A. adiantum-nigrum, which is present as a distinct morphological form. The characteristics of serpentine populations are carried over in cultivation, and while this species will grow in ordinary soil, it is less vigorous than when cultivated in a serpentine substrate (4). A. viride, which is considered an indicator of ultrabasic soils in Scandinavia, is only found on highly calcareous substrates in the United Kingdom (4). Equally noteworthy are the occurrences of several very rare interspecific Asplenium hybrids at UK serpentine sites [A. x alternifolia (A. septiontrionale x A. trichomanes), A. x contrei (A. septiontrionale x A. adiantum-nigrum), A. x murbeckii (A. septentrionale x A. ruta-muraria)]. It has been hypothesized that these may persist due to the lack of competition from other ferns and vascular plants (4). A. cuneifolium and the heliophile fern Nothalaena marantae are reported as abundant in shaded situations on serpentine rocks in southeastern France. The latter species is predominantly Mediterranean in its distribution and only occurs in isolated populations on serpentine substrates in central Europe, at the northern limits of its distribution (2). Other pteridophytes reported as typical for old mining areas in the UK, include A. ruta-muraria, A. trichomanes, and A. viride at alkaline sites, these being replaced by A. adiantum-nigrum and A. septiontrionale where conditions are more acidic (4). There are also characteristic pteridophyte communities associated with the metal-rich and acidic former tin mine settling ponds in southwestern England. Here, Anthyrium filix-femina is locally abundant and the horsetail Equisetum palustre dominates the pond margins (4). 5

GENETIC DIVERSITY

Studies on genetic variation in Mielichhoferia elongata, a species of copper moss with a widely disjunct distribution, have revealed some interesting facts. Most

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populations seem to possess a low degree of internal variability, but there is great variation between populations. Thirteen of the investigated populations seemed to consist of a single clone, but populations in Colorado possessed a tremendous resevoir of genetic variation. The genetic structure of M. elongata suggests repeated dispersal and founding of populations (31). Similar studies on a second copper moss Scopelophila cataractae, which has recently extended its range into Europe, show similar trends (32). European populations seem to be homogenous, as are populations from copper-enriched soils in the vicinity of Japanese temples, but other Asian and North American populations are very diverse. Long distance dispersal may account for the current world distribution of S. cataractae, it is probably native to both Asia and North America, but appears to be a recent immigrant to its western European sites. This species produces no sporophytes in its North American and European populations (33) but is fertile at its Asian sites. Studies on North American populations showed that 50% contained no gametangia and that male and female plants were never found in the same population. Plants from different populations showed extensive morphological variability when cultivated on differentially metal-contaminated substrates, indicative of genetic polymorphism. All populations grew best on the most contaminated substrate, with significant variations in growth being evident as substrate metal concentrations decreased (33). 6

CONSERVATION

Metal-enriched soils are under threat internationally. Areas of natural mineralization may be mined, while sites where soils have been polluted by metal extraction or processing are being increasingly rehabilitated to comply with national guidelines on acceptable concentrations of hazardous substances. Many isolated metalliferous outcrops in Zaire are being exploited, with no action taken to safeguard the unique plants that have evolved at such sites (A.J.M. Baker, pers. commun.). In the United Kingdom, old mine sites in former mining areas are being reclaimed, usually to some form of amenity grassland, while urban contaminated sites are frequently redeveloped in a manner that precludes the survival of any metallophytes. In the United Kingdom, several endemic and nationally endangered bryophytes are confined to a handful of former mine sites in Wales and southwestern England. Detailed conservation action plans have been drawn up to manage such locations to ensure the continued survival of the leafy liverworts Cephaloziella calyculata, C. integerrima, and C. nicholsonii and the mosses Ditrichum cornubiensis and Pohlia andalusica. The liverworts are threatened by landscaping or ‘‘tidying’’ of former mine sites as well as capping of old shafts and the associated building work. Ditrichum cornubiensis has more specialized requirements. This species is a colonist of bare, unshaded, copper-rich soils, such as track edges, eroding banks, and soil dug from ditches. The plant spreads by means of under-

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ground tubers and requires a succession of bare habitats to colonize. Any changes to the management of tracks or ditches that significantly reduce these types of habitat will seriously threaten the two extant populations of this global rarity (34). In North Wales, Ditrichum plumbicola is confined to lead mine spoil at approximately 20 locations in the Conwy Valley. It occurs on silty or peaty soils on stable waste tips, confined to sparsely vegetated areas and forming pure tufts on rough surfaces that may have been created by frost heave. It only occurs on acidic spoil tips and is absent where the tip has a pH ⬎6.5 (13). This species is clearly highly vulnerable to any management strategy that reduces bare ground or neutralizes soil acidity. In the same area, the cuprophile liverwort Cephaloziella masalongi, once frequently associated with copper mining sites, is now considered to be very rare, due to the cessation of mining and the gradual loss of suitable habitat (13). 7

CONCLUSIONS

Several species of bryophytes and pteridophytes are able to flourish on metalenriched soils and substrates. In some cases, species are solely confined to such sites, but other widespread species become dominant where substrate metal concentrations exert powerful selective pressures. Both plant groups have received significantly less attention than metallophytic vascular plants, leaving some important unanswered questions. There is a clear need for more detailed and controlled investigation of the copper mosses to establish the exact nature of the interaction between copper, sulfur, and pH in determining the ecological preferences of these species. In addition, there is also a pressing need for conservation of vulnerable metallophytic species, in a similar fashion to the detailed schemes proposed for endemic and endangered bryophytes of copper soils in the United Kingdom. REFERENCES 1. HT Shacklette. Copper mosses as indicators of metal concentrations. US Geol Surv Bull 1198G, 18pp, 1967. 2. P Duvigneaud. (1966) Note sur la biogeochimie des serpentines du Sud-Ouest de la France. Bull Soc Roy Bot Belg 93:271–329, 1966. 3. W Ernst. Schwermetallvegetation der Erde. Stuttgart: Fischer, 1974. 4. C Page. Ferns. London: Collins, 1988, pp. 270–278. 5. RR Brooks, F Malaisse. The Heavy Metal-Tolerant Flora of Southcentral Africa. Rotterdam: Balkema, 1995. 6. DHS Richardson. The Biology of Mosses. Oxford: Blackwell, 1981. 7. HA Hyde, AE Wade, SG Harrison. Welsh Ferns, ed. 4th Cardiff: National Museum of Wales, 1969.

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8. H Persson. Studies in ‘‘copper mosses’’. Hattori Bot Lab 17:1–18, 1956. 9. A Schatz. Speculations on the ecology and photosynthesis of the ‘‘copper mosses.’’ Bryologist 58:113–120, 1955. 10. A Sotiaux, Ph De Zuttere. Le genre Scopelophila (Mitt.) Lindb. en Europe. Cryptogamie Bryologie Lichenologie 8:95–108, 1987. 11. A Noguchi. On some mosses of Merceya with special reference to the variation and ecology. Kumamoto J Sci 2:239–257, 1956. 12. AJE Smith. The Moss Flora of Britain and Ireland. Cambridge, UK: Cambridge University Press, 1978. 13. MO Hill. A bryophyte flora of North Wales. J Bryol 15:377–491, 1988. 14. AJE Smith. The liverworts of Britain and Ireland. Cambridge, UK: Cambridge University Press, 1990. 15. A Empain. Heavy metals in bryophytes from Shaba province. In: RR Brooks, F Malaisse, eds. The Heavy Metal-Tolerant Flora of Southcentral Africa. Rotterdam: Balkema, 1985, pp 103–117. 16. HT Shacklette. Bryophytes associated with mineral deposits and solutions in Alaska US Geol Surv Bull 1198C 18pp, 1965. 17. DR Slingsby. The Keen of Hamar, Shetland—A long-term site-specific study of a classic serpentine site. In: AJM Baker, J Proctor, RD Reeves, eds. The Vegetation of Ultramafic (Serpentine) Soils. Andover, UK: Intercept, 1992, pp. 235–241. 18. DC Fraser. A syngenetic copper deposit of recent age. Econ Geol 56:961–962, 1961. 19. NW Lepp, D Salmon. A field study of the ecotoxicology of copper to bryophytes. Environ Pollut 106:153–156, 1999. 20. P Wilkins. Observations on the ecology of Mielichhoferia elongata and other ‘‘copper mosses’’ in the British Isles. Bryologist 80:175–181, 1977. 21. A Samecka-Cymerman, AJ Kempers. Bioaccumulation of heavy metals by aquatic macrophytes around Wroclaw, Poland. Ecotox Environ Safe 35:242–247, 1996. 22. H Wild. Geobotanical anomalies in Rhodesia. 3—The vegetation of nickel-bearing soils. Kirkia 7 Suppl. 1–62, 1970. 23. H Wild. Indigenous plants and chromium in Rhodesia. Kirkia 9:233–241, 1974. 24. H Wild. Geobotanical anomalies in Rhodesia. 4—The vegetation of arsenical soils. Kirkia 9:243–264, 1974. 25. H Wild. Arsenic-tolerant plant species established on arsenical mine dumps in Rhodesia. Kirkia 9:265–278, 1974. 26. JB Hall. Pteris vittata Linn. A gold mine fern in Ghana. Nigerian Field 35:1–9, 1970. 27. EH AmonooNeizer, D Nyamah, SB Bakiamoh. Mercury and arsenic pollution in soil and biological samples around the mining town of Obuasi, Ghana. Water, Air Soil Pollut 91:363–373, 1996. 28. H Wild. The flora of the Great Dyke of Southern Rhodesia with special reference to the serpentine soils. Kirkia 5:49–86, 1965. 29. WGB Jacobsen. A checklist and discussion of the flora of a portion of the Lomagundi district, Rhodesia. Kirkia 9:147–207, 1973. 30. JC Vogel, FJ Rumsey, JJ Schneller, SJ Russell, JS Holmes, JA Barrett, M Gibby. The origin, status and distribution of Asplenium presolanense spec. nov. (Aspleniaceae, Pteridophyta). Botanica Helv 108:269–288, 1998.

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31. AJ Shaw, RE Schnieder. Genetic biogeography of the rare copper moss Mielichhoferia elongata (Bryaceae). Am J Bot 82:8–17, 1995. 32. AJ Shaw. Genetic biogeography of the rare copper moss Scopelophila cataractae (Pottiaceae). Plant System Evol 197:43–58, 1995. 33. AJ Shaw. Population biology of the rare copper moss Scopelophila cataractae. Am J Bot 80:1034–1041, 1993. 34. Cornwall Biodiversity Initiative (2000). Metalliferous mines action plan. http:/ /www. wildelifetrust.org.uk/cornwall/wow/audit2/act aal.htm (accessed 31.01.2000).

7 Angiosperms (Asteraceae, Convolvulaceae, Fabaceae and Poaceae; other than Brassicaceae) Anna Siedlecka, Anna Tukendorf, Ewa Sko´rzyn´ska-Polit, Waldemar Maksymiec, Małgorzata Wo´jcik, Tadeusz Baszyn´ski, and Zbigniew Krupa Maria Curie-Skłodowska University, Lublin, Poland

Heavy metals are defined as metallic elements with atomic number higher than 20, but 12 of them are widely known as environmental pollutants due to their release by industry: Cd, Co, Cr, Cu, Fe, Hg, Mn, Mo, Ni, Pb, Sn, and Zn. Within this group Cd, Cu, Fe, Hg, Ni, Pb, and Zn are usually considered as potentially hazardous for plants and animals, despite the fact that most of them (Cu, Fe, Ni, and Zn) are known as essential elements necessary for plant and animal metabolism (1–3). In general, biodiversity in plant response to heavy metals will be discussed in this chapter. Nonetheless, some particularly important responses to Al—nonheavy, but well known as a very toxic metal—are also described. Some plants, like Arabidopsis sp. (a model organism in plant molecular biology), Thlaspi sp. (hyperaccumulator), or such popular experimental material as Brassica sp., Lactuca sp., or Raphanus sp. belong to the Brassicaceae 171

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family, and their responses to metals are described in detail by Di Toppi et al. in Chapter 8 of this book. 1 1.1

METALS UPTAKE Soil Factors

Biodiversity in plant response to heavy metal stress results from a number of factors: the metal itself, the plant species, and even the plant growth stage. The content of heavy metals in natural, uncontaminated soils is very different and depends on a number of factors, such as lithosphere (base rock), soil age, soil type, and covering flora. In general, heavy metals can be accumulated in higher amounts in the sorption complex of soils rich in organic matter, but their release to soil solution is much slower than in mineral soils due to high affinity of soil organic compounds to heavy metals (1). Heavy metals solubility in soil and thus their availability to plants is controlled also by soil pH, type of mineral colloids, and many other important factors, such as microbial activity, redox potential, and aeration (1,2) (Fig. 1). Some of these factors act in several ways; an increase

FIG. 1 Soil and root factors modifying heavy metals uptake from soil.

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of Ca2⫹ content in soil (liming) decreases heavy metals uptake by roots due to physiological antagonism of Ca2⫹ /heavy metal ions, but heavy metals uptake is decreased even more by increased heavy metals retention by soil colloids (1) (Fig. 1). Availability of free ions has been widely considered as a primary factor of metallic nutrients uptake by plants. There is evidence now that this relationship is not so obvious (4). Free ion concentration dependent uptake of divalent cations is easier than when metallic nutrients are present in chelated form (5) (Fig. 1). Nonetheless, if the quantity of the metal-complexing ligand is limited (the case of soil solution ligands, which often are of low affinity and low selectivity in respect to individual elements) or the ligand is very diluted, chelation may have no influence or may even enhance a divalent cation uptake (4). However, chelation has a very stimulating effect on root uptake and translocation to shoots of trivalent cations, like Cr, Ga, or In (5) (Fig. 1). It is also not possible to select the most important soil factors influencing metals availability to plants because their importance can vary among elements; for Zn it seems to be a Zn-buffering mechanism in the soil, whereas for Fe and Mn it is soil pH and soil redox state (2,6,7). 1.2

Plant Factors

Almost all heavy metals can be taken up by plants in two ways, which are mostly concentration-dependent (2,3,7–12): Nonmetabolic uptake by energy-independent mechanisms. Intact membranes are effective barriers for ions and uncharged molecules, but when solutes are more concentrated at the one side of the membrane they can diffuse down the concentration gradient with the aid of membrane carriers or even aqueous pores. This transport is known as passive and takes place when heavy metal is present in the root environment at a high amount. It may be also stimulated by lowering the free ion level in cytoplasm due to its incorporation into organic structures, deposition in some cell compartments, transport to other cells, or bounding to charged groups. This process often takes place in meristematic tissue of root tips. Metabolic uptake by energy-dependent mechanisms. This active uptake is involved in taking up ions against their concentration gradient. In this mechanism, a proton motive force (ATP-driven H⫹ pumps) creates pH and electropotential gradient which stimulate ion passing to the cell through selective ion channels or carriers. Higher plants have sophisticated and specific mechanisms for Fe uptake. Iron, despite being widely spread in the lithosphere, usually predominates in soils in

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nonsoluble and oxidized forms, and plants have developed two strategies of its uptake involving two main problems: availability and necessity of its presence in plants in reduced form (1,2,13–18). Strategy I, typical for most higher plants (all dicotyledonous and monocotyledonous except Graminae), is a three-step, complicated process. The first step—mobilization—means release of protons, organic acids, and phenolic compounds to the soil resulting in evolving free Fe3⫹ from the soil complex and its chelation by phenolics to keep it in solubilized form. The next step is reduction by root membrane reductases—constitutive proteins whose activity increases even by several orders of magnitude in Fe deficiency conditions. Reduction of iron from its ferric to ferrous form takes place in the rhizosphere, outside roots. The last step is ferrous ion uptake by roots, mainly through plasmalemma of special transfer cells. The number of these cells increases significantly in Fe deficiency conditions. Strategy II, present only among Graminae, is much simpler than strategy I. Roots of grasses release to soil specific chelating substances called phytosiderophores, which are mainly derivatives of mugineic acid synthesized from nicotianamine—a nonprotein amino acid. They extract and chelate ferric ions from soil sorption complex; the whole complex with oxidized Fe ions is taken up by plant. Ferric ions are released and their immediate reduction takes place inside root cells. Strategy II is more effective and more resistant to unfavorable environmental factors than strategy I because two important steps (Fe release from chelate and Fe reduction) are shifted from the rhizosphere into homeostatic conditions inside living root cells. 1.3

Interactions in Uptake

For plants of both strategies heavy metals toxicity results in decreased Fe uptake and Fe deficiency despite a good availability of this essential element. This effect is so strong that external symptoms of heavy metal toxicity and Fe deficiency are very similar for a number of elements (2,19–22). Inhibition or disturbances in functioning of root plasmalemma ATPase and oxidoreductase as well as young root tip damage and increased root Fe immobilization belong to the main mechanisms of Fe deficiency induction in strategy I plants (2,3,9,22–26). Strategy II plants are less sensitive to heavy metals–induced Fe deficiency due to better efficiency of this strategy, as described above. Nonetheless, inhibition of mugineic acid–ferric complex uptake in barley roots by Cu, Zn, and Co was also reported (27).

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Toxic influence of heavy metals is connected not only with inducing Fe deficiency but with numerous disturbances in plant mineral composition, resulting in complicated effects both of indirect and direct nature, leading to changes in plant metabolism (2,3,22–26,28–30). In general, relationships between toxic metals and other nutrients, including heavy metals essential for plant growth and development, can be divided into three categories (3,31) (Fig. 2): Addition—when there are no effects for plant growth and metabolism of introducing yet another element to the environment Antagonism—when the introduced element is of beneficial influence for plants Synergism—when the new element results in multiplication of plant stress symptoms

FIG. 2 Plant responses to combination of metals in the growth medium. Different arrows show direction of relationship documented: ↔, same relationship occurs for both metals; →, one metal influences another one. (Data taken from Refs. 10, 31–35.)

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PHOTOSYNTHESIS

The sensitivity of higher plants to heavy metals stress is revealed largely by its interference in functioning of the photosynthetic apparatus (36, 37 and references therein). Alterations of the chloroplast function and structure in plants exposed to heavy metals depend on the individual heavy metal, its concentration in the plant growth medium, uptake to plant organs, and on many mechanisms involved in detoxification processes after the metal has entered the plant as well as its final concentration found in chloroplasts. Plant species, their genetic and individual features, duration of heavy metals action, the age and physiological condition of plants during the treatment are also greatly significant. The photosynthetic apparatus seems to be affected by a number of direct and indirect actions of the metal ions resulting in functional and structural disorders. 2.1

Photosynthetic Pigments

One of the toxic effects resulting from excess supply of heavy metals is a change in chloroplast pigments. Reduced accumulation of chlorophyll as an effect of the action of heavy metals such as Cd, Cu, Hg, Pb, Mn, Ni, and Zn in many plant species has been well documented in both in vitro and in vivo experiments (37– 45). Chlorosis, the most apparent symptom of heavy metals effect in leaves, may result from inhibition of chlorophyll synthesis caused by reaction with constituent biosynthetic enzymes as well as chlorophyll degradation. Another reason of chlorosis may be a strong interaction between heavy metals and Fe (34, 37 and references therein, 40, 46, 47). Heavy metals–induced Fe deficiency obviously affects, among others things, photochemical activity of thylakoid membranes, including all Fe-containing complexes of the photosynthetic electron transport chain (3, 18 and references therein). The opinion that heavy metals, such as Cd, Hg, or Pb, interfere with chlorophyll biosynthesis by inhibition of δ-aminolevulinic acid (ALA) formation, ALA dehydratase activity as well as protochlorophyllide photoreduction is based on experiments with short-term action of the metal in detached leaves or their segments as well as in whole plants (48–55). It was recently postulated that Cd does not affect chlorophyll biosynthesis itself but interferes with chlorophyll integration into stable chlorophyll-protein complexes of thylakoid membranes required for normal protosystem II (PSII) activity (43). The metal, reducing chlorophyll content in leaves, also influenced the expression and assembly of chlorophyll-binding proteins. A strong reduction of all chlorophyll-containing complexes in cucumber under Cd treatment was found in the order of PSI ⬎ LHCII ⬎ PSII core and to a lesser extent under Pb stress, where LHCII appeared somewhat more sensitive than PSI (45,47). Cadmium effects on disturbances of LHCII oligomerization process correlated with

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the level of trans-16:1 fatty acid in phosphatidylglycerol were hypothesized (56– 58). Most recently, Cd affected LHCII accumulation by drastically reducing the steady-state level of Lhcb transcripts was demonstrated on the basis of hybridization analysis (59). Catabolism of chlorophyll in heavy metal–treated plants was also sporadically considered. The expected mechanism of chlorophyll degradation related to stimulation of chlorophyllase activity in heavy metal–stressed plants has not been univocally documented. An increase in hydrolytic activity of chlorophyllase in rice leaves was induced by Hg, Zn, and Cu (60), in contrast to its decrease with increasing Cd concentration found in primary leaves of barley (61). Involvement of Cd-induced lipid peroxidation in chlorophyll degradation based on a positive relationship between increase of lipoxygenase activity and decrease in chlorophyll content was recently proposed (62). The participation of galactolipase, whose activity is enhanced in Cd-treated plants, was also taken into consideration in diminished chlorophyll accumulation (63). Chlorophyll accumulation in plants stressed by heavy metals depends on their sensitivity or tolerance to metal action. A differentiated step of chlorophyll synthesis inhibition was observed in sensitive (significant chlorophyll decrease) and tolerant (almost unchanged chlorophyll level) population of spinach as a response to Cu toxicity (64,65). Also, in a Cu-tolerant population of Silene cucubalus chlorophyll content was not affected by Cu, whereas sensitive plants became chlorotic (66). These toxic effects of the metal were confirmed later when chlorosis in a Cu-sensitive population of Minuartia hirsuta and permanent green leaves in a Cu-tolerant plant were demonstrated (67). In wheat cultivars, which differ in their tolerance to Mn, chlorophyll concentration declined under Mn treatment. However, a higher pigment concentration was maintained in tolerant than sensitive cultivars (68). Also, three genotypes of Phaseolus vulgaris with contrasting tolerance to Mn toxicity showed increased chlorosis in the more sensitive genotype (69). Differentiated resistance of plants to heavy metals, resulting in changes of the plastid pigment content, depends on the developmental stage of the plants in the time of treatment. In runner bean plants treated with Cu and Cd at the early growth stage the chlorophyll content increased on the leaf area basis, while chlorosis was observed in plants treated at the final stage of primary leaf growth (70– 74). A reduced leaf area, an increased leaf density, smaller palisade parenchyma cells, and reduced mesophyll cell volume in young leaves may suggest a lower sensitivity of the pigment to Cd and excess Cu than processes of leaf growth. Chlorophyll content depending on the developmental stage was observed in leaf sections of different ages of monocotyledonous plants, such as maize (75) and rye (76). The relationship between the developmental step of pigeon pea exposed to Cd and Ni and chlorophyll content was indicated by almost unchanged pigment

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content in plants when treated with heavy metals at the vegetative and flowering step. However, drastic reduction in chlorophyll content, greater in Cd- than in Ni-treated plants, was observed at the pod-filling stage (77,78). 2.2

Chloroplast Ultrastructure

Changes in chlorophyll accumulation are usually accompanied by heavy metal– induced drastic ultrastructural damages of the chloroplast fine structure, mainly at the thylakoid level (46,64,65,75,79–96). During a long-term action of heavy metals, the decomposition of the chloroplast architecture, such as degradation of grana stacks and stroma lamellae, increase in the number and size of plastoglobuli, and occurrence of intrathylakoidal inclusions, is the most frequently observed symptom in heavy metal–stressed plants. In damages of chloroplasts as a response of plant to heavy metals, the developmental stage of plants during the treatment as well as plant sensitivity to their action seem to be of great importance. Small or no changes in chloroplast ultrastructure of runner bean plants treated with Cu and Cd at their early stage but great alterations at the end of growth stage have been observed (87,90). The tissue sections at different stages of maturity along secondary maize leaf revealed that at high Cd concentrations ultrastructural alterations of chloroplasts were more distinct in mature tissue of both the mesophyll and the bundle sheet (75). These observations can suggest dicots and monocots to be more resistant to heavy metals at the early growth stage. Nevertheless, the tolerance of monocots to heavy metals is generally higher than that of dicots and requires rather a long exposure time to heavy metals. Heavy metals affect chloroplast ultrastructure to a greater extent in sensitive than tolerant wheat cultivars (82) and spinach populations (64,65). Delayed etioplast transformation into chloroplast caused by Pb (94) and Cd (85) was also reported. 2.3

Thylakoid Membrane Polypeptides and Acyl Lipids

Disruption of the chloroplast ultrastructure in heavy metal–stressed plants seems to be related to disorganization of protein and acyl lipid composition of thylakoid membranes. Heavy metal–affected damage to the structure and composition of the thylakoid membranes induces changes in lipid matrix, while concomitant release of fatty acids probably leads to dissociation of thylakoid polypeptide components. Decomposition in chlorophyll-protein constituents of PSII was found in older spinach and bean plants exposed to Cu (65,97). A lower accumulation of some extrinsic polypeptides of PSII particles is age-dependent as shown in runner bean plants treated with Cd and excess Cu (70,74,89). In older chlorotic leaves it concerns a high decrease in the polypeptide content both of the oxygen evolving complex (OEC) and PSII core antenna. In leaves treated with the metal at a

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younger age only 17- or 17- and 23-kDa polypeptides of OEC changed noticeably. Polypeptide release from thylakoid membranes in rice exposed to excess Cu was also noted (98). These data correspond well to changes in PSII activity confirming the donor side of PSII as a most sensitive site of heavy metal action along the photosynthetic electron transport chain proposed earlier (37 and references therein). Experiments in vitro are in agreement with this statement. Disorganization of the donor side of PSII is indicated by the release of OEC polypeptides from thylakoid membranes of spinach (99) and Vigna unguiculata (100) incubated with Cd. Similar conclusions resulted from significant depletion of 17- and 23kDa polypeptide and partially of the 33-kDa polypeptide in both Pb- and Znincubated spinach chloroplasts (101). Heavy metal–induced damages in thylakoid membranes were visualized also by alteration in the lipid fraction of the membrane. The changes in the content and composition of acyl lipids forming the lipid environment of photosystem complexes were confirmed by several authors. A decrease of acyl lipids and changes in their fatty acid composition in the chloroplast membranes were observed after exposure of Phaseolus vulgaris and Zea mays to Pb (102–104), Lycopersicon esculentum, Raphanus sativus, and Phaseolus coccineus to Cd (56,58,63,74), Spinacia oleracea, Phaseolus coccineus, and Triticum durum to excess Cu (70,95,97), as well as Pinus banksiana to V and Ni (105). An alteration in the lipid environment around PSII was taken into account as the cause of the loss of PSII activity following Cd treatment (37,57,58). Some of these changes that implicate photosynthetic activity in chloroplasts depended on the time of metal action, its concentration and application to the nutrient solution, and the age/growth stage of plants. The destruction of the thylakoid membrane shown in spinach chloroplasts was interpreted as a result of lipid peroxidation mediated by excess Cu (106). Significant changes in acyl lipid correlated with a distinct loss of core antenna PSII polypeptides and OEC subunits mentioned above characterize spinach and bean plants exposed to Cu and Cd by the end of the intensive growth stage of primary leaves (74,97). 2.4

Photosynthetic Electron Transport

Photochemical reactions of PSII have been shown to be more susceptible to heavy metal stress than those of PSI. On the basis of in vitro and in vivo studies, the PSII reaction center and PSII electron transport were recognized as strongly affected by Cd (37 and references therein, 107) (Fig. 3). For a long time PSI was considered as relatively resistant to Cd. In earlier in vivo studies, only a small decrease in PSI activity in Cd-treated plants was found (81,85). Later on, at higher concentrations of the metal, a significant inhibition of both PSII and PSI in chloroplasts of Cd- and Pb-treated clover and lucerne seedlings was reported (108) and

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FIG. 3 Toxic influence of metals on the photosynthetic apparatus of C3 higher plants. Abbreviations: CA, carbonic anhydrase; CF0 , chloroplast coupling factor 0; CF1, chloroplast coupling factor 1; cyt, cytochrome; Fd, ferredoxin; FDPase, fructose-1,6-bisphosphatase; FNR, ferredoxin-NADP⫹ oxidoreductase; GAPDH, NADPH-dependent glyceraldehyde-3-phosphate dehydrogenase; LHC I, light-harvesting chlorophyll a/b protein complex I; LHC II, lightharvesting chlorophyll a/b protein complex II; OEC, oxygen evolving complex; PC, plastocyanin; PCR, photosynthetic carbon reduction cycle (Calvin cycle); PGA, 3-phosphoglyceric acid; PQ, plastoquinone; PR kinase, ribulose-5-phosphate kinase (phosphoribulokinase); PS I, photosystem I; PS II, photosystem II; RA, Rubisco activase; Rubisco, ribulose-1,5-bisphosphate carboxylase/oxygenase; SdBPase, sedoheptulose-1,7-bisphosphatase; TM, thylakoid membrane; TPI, triosephosphate isomerase. (Data taken from Refs. 18,35,37, 39,52,77,79,115,130–138). Cu influence on Rubisco and CA in Phaseolus vulgaris plants—Siedlecka and Krupa (unpublished data).

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also confirmed in chloroplasts of Cd-exposed runner bean plants (111). At the same time, new data showing that application of a low, non toxic level of Cd to wheat plants modifies the primary photochemistry of PSI as that of PSII were obtained, although its sensitivity to Cd was lower than that of PSII (112). A similar conclusion was drawn from PSI activity inhibition in respect of Cd-Fe interaction in bean plants (18 and references therein, 35) (Fig. 3). Recently, it was demonstrated that in maize seedlings Cd strongly interacts with Fe leading to Fe deficiency, thus decreasing the level of ferredoxin and markedly affecting PSI activity on its reducing side (24). The sensitivity of the PSI reducing side was suggested in earlier in vitro studies on spinach chloroplasts incubated with Cu, showing not only PSII disorder on its donor side but also PSI inhibition due to interaction with ferredoxin (111). However, the PSI oxidizing side affected by heavy metals was recognized in vivo when a decrease in plastocyanin content in chloroplast of rice exposed to excess Cu (112) as well as in Ocimum basilicum plants treated with Ni and, to a lower extent, Zn was shown (113) (Fig. 3). Decrease in ferredoxin content in chloroplasts of O. basilicum indicates that both sides of PSI are Ni-sensitive. The sensitivity of both PSII and PSI to Cu was also found in Cu-treated intolerant spinach (65) and ecotype of Thlaspi ochroleucum (42), but changes in PSI after exposition to Cu were smaller than those occurring in PSII. PSII, essential for photosynthetic regulation, is strongly affected by heavy metals at different target sites (37, 39, 48, 114 and references therein) (Fig. 3). As shown in in vitro studies, the water-oxidizing system of PSII was the most affected, but OEC was postulated as the primary target of heavy metals toxicity. This was supported in in vivo examinations of photosynthetic activities inhibition in plants under Cd, Pb, and excess Cu stress (81,85,108). The replacement of Mn ions in the OEC by Cd (81) and Cd- or Cu-induced alteration in the lipid composition of the thylakoid membrane resulting in release of fatty acids causing OEC subunits disturbance were taken into consideration as the reason for the loss of PSII activity. Inhibition of photosynthetic electron transport by Cu, mainly at the PSII before the DPC (1,5-diphenylcarbohydrazide) donor site, was shown in rice chloroplasts (115) (Fig. 3). The PSII electron transport on both the donor and acceptor side influenced by heavy metals, like Cu, was concluded elsewhere (67,116). The antenna chlorophyll a molecule of PSII was also proposed as a primary site of Cu inhibition (86). The inhibition site for Co, Ni, and Zn in chloroplasts isolated from pea leaves was located on the PSII acceptor side at the level of the secondary quinone acceptor, QB (117) (Fig. 3). This was also indicated by increase in QA reduction state and thermal energy dissipation due to excess Mn limiting energy utilization in the Calvin cycle found in white birch leaves (118). The acceptor side as a site of Cu action in chloroplasts of runner bean plants was also confirmed in in vivo studies (70,71).

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The toxic effect of heavy metals on the photosynthetic processes in leaves of developing and mature plants seems to be different. In hydroponically cultivated runner bean plants supplied with excess Cu various responses depending on the growth stage at which the metal was added to the nutrient medium were noticed (70,71). In plants treated with Cu at the young development stage, the inhibitory effect on the acceptor side of PSII, (due to induced inhibition of the Calvin cycle) and down-regulation of the electron transport was suggested in the absence of changes in the primary PSII photochemistry. In plants treated at a more developed stage a low PSII activity may result from alteration in the donor side and PSII reaction center as well as in the acceptor side of PSII. In addition, the effect of Cd on PSII activity was shown to be considerably modified by the developmental stage of plants during treatment (73,74). In runner bean plants treated with Cd at an early growth stage of primary leaves no disturbances of PSII efficiency and a small increase in heat dissipation of excitation energy were observed. Cadmium treatment at the end of the growth stage affected PSII photochemistry, electron transport, and dark reactions correlated with degradation of thylakoid membranes. The mechanism of heavy metals action on the photosynthetic apparatus can be influenced by the level of Ca2⫹ accumulation in plants. Enhancement of Cd toxic effect, including primary photochemistry of PSII, in runner bean plants grown in Ca2⫹-deficient medium was documented as well as reduction of Cd toxicity in Ca2⫹ excess (119). Similarly, in young runner bean plants Ca2⫹ deficiency increased the toxic effect of excess Cu on the photosynthetic apparatus (120–122). However, in older plants such an effect was observed at increased Ca2⫹ accumulation. It is hypothesized that in the first case Cu can substitute weakly bound Ca2⫹ present in OEC and/or coupling factor (123–125); in the other case, Cu through Ca2⫹ accumulation develops senescence processes (126 and references therein). A mechanism of partial prevention of Cu-induced PSII inhibition by Ca2⫹ was postulated for PSII particles depleted of OEC subunits (127). In heavy metal–induced degradation of the photosynthetic apparatus, disturbances of the turnover of D1 reaction center polypeptide of PSII has been signalized (128 and references therein). In bean plants, after exposure to excess Cu, the equilibrium between photoinhibition and repair resulted from increase in the quantum yield of photoinhibition (116). Inhibition of the turnover of D1 protein in plants under Cd stress depending on the plant species and the time of Cd exposure was also postulated (129). 2.5 2.5.1

Photosynthetic Carbon Assimilation C3 Plants

According to Weigel’s observations (139,140), inhibition of the dark reactions of photosynthesis by Cd precedes that of the photosynthetic electron transport.

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Numerous data obtained from chlorophyll a fluorescence induction kinetics measurements were recently published, leading to the conclusion that the Calvin cycle is probably the primary target of heavy metals toxicity (37 and references therein). Changes in the primary photochemical reactions and electron transport seem to be secondary effects, at least for young plants. The proposed mechanism of heavy metals toxicity concerns limitation of ATP and NADPH consumption in the Calvin cycle causing an increase in proton gradient across thylakoid membrane, finally leading to down-regulation of both PSII photochemistry and linear electron transport (see 37 and references therein, 141) (Fig. 3). Similar mechanism of heavy metals toxicity on the photosystems in stressed plants was also proposed in regard to the action of Ni, Cu, Pb, Fe, and Cd (35,44,47,70,74,82, 119,141,143). Weigel concluded that further steps of the Calvin cycle—reduction and regeneration—are the most sensitive targets of Cd toxicity (139,140). Inhibition of some important enzymes of these phases of the Calvin cycle was confirmed in Cajanus cajan, Phaseolus vulgaris, and Triticum aestivum in the presence of some heavy metals (52,77,132,134), (Fig. 3). Nonetheless, based on current information, Rubisco and carboxylation process carried out by this enzyme seem to be not only the most important but also the most sensitive step in plant response to heavy metals (Fig. 3). Rubisco is the most abundant protein in chloroplasts (more than 60% of total leaf protein) and its activity depends on next two enzymes present in excessive amounts in stroma: Rubisco activase (RA), 5% of total leaf protein, and carbonic anhydrase (CA), 2% of total leaf protein (137,144,145) (Fig. 3). Rubisco activase regulates correct conformation of Rubisco-active center. The pool of Rubisco reaction centers that are activated and have proper conformation to act as enzyme is called the Rubisco activation state (RAS) and is expressed as a percentage of all Rubisco-active centers. Usually RAS is at the level of 60– 70% (137,144). Depending on the Cd concentration, two different mechanisms of Rubisco activity regulation are proposed (18, 137 and references therein): At low Cd level, CA activation is enough to maintain full Rubisco activity, even with accelerated Rubisco light activation at the beginning of light day period (136,138). At higher Cd concentrations, inhibition of CA occurs and ATP-dependent (probably RA-dependent) mechanism results in increase of RAS up to 100%. However, this mechanism is not as efficient as the CA-dependent one, so despite this effort Rubisco activity decreases (136,138). An interesting phenomenon is that the above-described mechanism of CAdependent maintenance of Rubisco activity seems to operate only at moderate Cd stress. Other investigated metals, i.e., Fe (136) and Cu (Siedlecka and Krupa, unpublished data), even at low toxic concentrations caused substantial decrease in CA and Rubisco activities (Fig. 3). It is also confirmed that monocotyledonous plants showing C3 metabolism, like rye or wheat, are less sensitive to heavy

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metals stress than dicotyledonous plants, also with respect to the dark phase of photosynthesis (30,39). 2.5.1

C4 Plants

Among C4 plants the primary carboxylation process, catalyzed by phosphoenolopyruvate carboxylase (PEP-carboxylase), is the most sensitive target of heavy metals toxicity (37, 39, 146 and references therein). In Zea mays PEP-carboxylase appeared to be very sensitive to Cd, Cu, Pn, and Zn (130). On the other hand, the C4 plants are well known as less sensitive to heavy metals stress than C3 plants, including C3 monocots. The reason is probably two-step carboxylation: the first step is PEP-carboxylase dependent and sensitive to stress conditions, but as long as this carboxylation occurs, the Rubisco-dependent secondary carboxylation is maintained by the usual CO2-concentrating procedure, resulting in efficient metabolism despite stress conditions. 3 3.1

RESPIRATORY PROCESSES Photorespiration

It is well known that heavy metals stress may cause a shift in Rubisco activity from carboxylation to oxygenation (39; for review, see 137). On the other side, the protective role of photorespiration to the photosynthetic apparatus in environmental stress conditions was also recently reviewed (147). Peroxysomal enzymes of Phaseolus vulgaris were inhibited by toxic Zn concentrations (133). Interesting changes in peroxysomal NADP-hydroxypyruvate reductase (HPR1) were recently observed in Phaseolus vulgaris, Secale cereale, and Arabidopsis thaliana in the presence of Cd. At low Cd concentration an increase in the amount and activity of this enzyme was observed, while a high Cd concentration resulted in decreased HPR1 activity (148). For rye plants changes in HPR1 activity were the smallest, confirming a higher resistance of monocots to heavy metals stress conditions. 3.2

Respiration

So far much less attention has been paid to effects of heavy metals stress on respiration than on photosynthesis. Mitochondria are known to be much more resistant to heavy metals than chloroplasts, and they remain undisturbed even at Cd high concentrations (149). Nonetheless, it is obvious that respiration may be affected by heavy metals in both direct and indirect ways. Direct inhibition results from inhibition of enzyme activities (for review, see 48). Indirect inhibition comes from decreased gas exchange due to limited stomata conductance (146). Insufficient production of ATP and NADPH by heavy metals–damaged chloroplasts causes energy imbalance in plant cells. If the stress is not too strong the

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plant tends to adapt by increase in respiration, which was confirmed by detected increase in some Krebs cycle enzymes content and/or activity in Silene italica and Glycine max under Cd, Pn, and Ni toxicity (150; for review, see 48). 4

ROLE OF BIOCHEMICAL MEDIATORS IN DIFFERENTIATION OF METABOLIC RESPONSE OF PLANTS TO HEAVY METALS STRESS

Several groups of substances, such as Ca2⫹ ions, systemin, salicylic acid (SA), ethylene, and jasmonians (JAs), act as different stress signals within the plant organs. These signal intermediates have various effects on metabolic processes depending on plant species, growth properties, and individual character and strength of the signal factor (151–156). Copper ions, involved in many oxidoreductive processes, are known as inducing ethylene synthesis in spinach and Scenedesmus plants (105). A high level of ethylene, resulting from inhibition of photosystems, can increase senescence processes observed in dicots (spinach and bean) and monocots (wheat) plants treated with Cu2⫹ at the final growth stages (70–72,120) or after a longer exposure to the metal (83,97). However, in bean plants increased lipoxygenase activity caused by Cu2⫹ (and also Zn2⫹) ions, accompanied by increased ethylene production, was observed (157). It indicates that Cu2⫹ can also induce free radical processes in cell membrane components directly (158–162) and in consequence triggers the JAs signal pathway (126). In many cases JAs stimulate ethylene biosynthesis (163–165); however, its inhibition has also been found in cocklebur seeds (166). More likely JAs stimulate ethylene synthesis (or increase in ethylene sensitivity) and simultaneously accelerate senescence processes, particularly after a wounding stress or in older plants. At present, data from one study indicate that in rice and Arabidopsis Cd and Cu ions can induce octadecenoic pathway (Table 1). However, in Rauvolfia serpentinea culture cells this effect was not observed and the ethylene content was decreased (182). Phaseolus coccineus is also sensitive to excess Cu or Cd, but this sensitivity depends on the plant growth stage (119,126). It indicates that Cd and Cu can influence this plant species through JAs-independent pathway. The induction of ethylene synthesis by Fe2⫹ ions varied in Oryza sativa plants (Table 1). In this case, plants previously exposed to wounding stress showed increased sensitivity of ethylene production to excess metal (179). Among the investigated metals only Co ions showed inhibited ethylene synthesis (180,181), whereas Ag affected ethylene action (183). These results indicate that such factors as growth conditions, stress intensity (186), and probably plant species may change the ways of heavy metals action. Many heavy metals may elicit proteins similar to PR (pathogenesis-related) proteins, which potentially may also be induced by signal mediators such as JAs or SA. They are Cu, Cd, Zn, Al, Fe, La, Ag, and Ga ions investigated in Triticum

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TABLE 1 Heavy Metals Stress and Expression of Signaling Pathways in Higher Plants. Physiological effect [Ca2⫹]c concentration (Ca-dependent signalling): Increased Decreased

Expression of stress-responsive gene: Induced

Not induced Ethylene content: Increased

Decreased

Metal

Al Al

Cu, Al, Cd, Co Cd Cu, Cd, Zn, Al, Fe, Zn, Ga, La Ag, Pb, Cu, Cd Cu, Cd, Zn Cu Cu Cu, Zn Fe2⫹ Co Cd, Cu Ag

Octadecenoic pathway: Induced Not induced

Cd, Cu Cu Cd, Cu

Plant

Refs.

Hordeum vulgare Triticum aestivum Amaranthus tricolor Triticum aestivum Tobacco cells

167 168 169 170 171

Hordeum vulgare Datura inoxa Triticum aestivum

172 173 174 175 176 177 178

Lupinus luteus Mimulus guttatus Glycine max Spinacia oleracea Phaseolus vulgaris Oryza sativa Cucumis sativus Nicotiana tabacum Rauvolfia serpentina Cucumis sativus, Lycopersicon aesculentum

105 157 179 180 181 182 183

Arabidopsis thaliana Oryza sativa Rauvolfia serpentinea

184 185 182

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TABLE 2 Relationships Between Some Plant Hormones and Heavy Metals Toxicity Plant hormone GA3 ABA

Auxin

Effect

Plant

Ref.

Partially reverses the effect of Cd or Ni Enhances plant growth inhibition by Cd or Ni; does not affect the influence of Al on growth Similarly to Cd, Cu, and Ag, induces accumulation of the same heat shock–like mRNA class

Oryza sativa

192

Oryza sativa

192,193

Glycine max

194

aestivum (174,175,187), Datura innoxia (173), and Lupinus luteus (176). However, in Mimulus guttatus and Glycine max, Cu, Cd, and Zn did not induce such a phenomenon (177,178). Such differences may be connected with plant species or with methodological incoherence, and require further studies. The cytoplasmic free Ca2⫹ ([Ca2⫹]c) is probably related to Ca-dependent signal transduction pathway. Aluminum ions, a very potent plant growth inhibitor, affected this pathway when compared with heavy metals (171) (Table 1). Depending on the plant species (Table 1), external Ca content in the medium, as well as plant organs (167), Al ions affected cell metabolism, cytoskeleton structure, cytoplasmic pH through an increase or decrease in [Ca2⫹]c, and in consequence inhibited growth processes (188,189). Some heavy metals also strongly affected Ca content in the plant tissues (67,119–122,190,191). Minor data about heavy metals–plant hormone interactions, except those of ethylene, are summarized in Table 2. They indicate that so far only in the case of ABA (in many cases similar in its action to JAs) and auxin has the response of the plant hormones to heavy metals stress been shown. 5 5.1

BIODIVERSITY AND OXIDATIVE STRESS Reactive Oxygen Species

Oxygen is evolved during photosynthesis and its consumption, more than 85% used by the cell, occurs in mitochondrial respiration. Molecular oxygen in its ground state is unreactive, but it can be reduced by four electrons to H2O. During O2 reduction, intermediates such as superoxide radical (O2⫺• ), hydrogen peroxide (H2O2), and hydroxyl radical (OH•), known as reactive oxygen species (ROS),

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can be generated. Singlet oxygen (1O2) is another kind of ROS, which is formed due to excitation energy transfer from chlorophyll to O2. ROS are capable of reacting at the site of their generation (or some of them can diffuse to another part of the cell) with all cellular compounds, causing lipid peroxidation, protein damages, and DNA mutation, forming chain reactions of free radicals. Generation of free radicals leads to peroxidative damage of membranes and often to irrepairable metabolic dysfunction and cell death. To prevent such situations, an active antioxidant system is present in plant cells, that keeps ROS formation under control. Antioxidant system includes such enzymes as catalase (CAT), peroxidases (POXs), superoxide dismutases (SODs), glutathione reductase (RG), and low molecular nonenzymatic compounds. They act as scavengers, first of all O2⫺• , H2O2, and 1O2 to prevent generation of the most toxic ROS, i.e., the hydroxyl radical (OH•). ROS are also formed in higher plants treated with heavy metals, but unfortunately this has been shown only in a few papers, so that it is difficult to discuss the biodiversity in ROS formation under this kind of stress. Some experiments were carried out on isolated chloroplasts or detached plant organs using the infiltration technique, but only a few were done on heavy metals treatment plants growing in nutrient solutions. Yruela et al. (195) showed in vitro OH• formation in chloroplasts membranes of Beta vulgaris treated with Cu. Lead (only at sublethal concentrations) caused an increase in the total pool of free radicals in lupine roots (199). An increase of H2O2 was found in leaves Pisum sativum infiltrated by Cd (197) and in leaves of Phaseolus vulgaris treated with Zn (198). A short-term treatment of potato tuber discs with CdCl2 increased the concentrations of H2O2 and O2⫺• . Moreover, in the susceptible potato species the increased H2O2 level lasted longer than in the tolerant one (199). Age-dependent level of ROS was observed in primary leaves of Phaseolus coccineus grown in the nutrient solution containing Cd. The increase in O2⫺• level was observed in Cd-treated young plants (Sko´rzyn´ska-Polit et al., unpublished). 5.2 5.2.1

Enzymes of Oxidative Stress Superoxide Dismutase

Superoxide radical can spontaneously dismute to H2O2 and O2 in reaction accelerated by SOD. Superoxide dismutases are the family of enzymes containing different metals in their active sites. SODs containing Cu and Zn (Cu/Zn-SOD) are generally found in cytosol of eukaryotic cells and chloroplasts, while those containing Mn (Mn-SOD) are found in the matrix of mitochondria and in prokaryotes (200). Iron-containing SODs (Fe-SOD) are mainly present in prokaryotic organisms and some eukaryotic algae. They have been also found in some families of higher plants: Gingkoaceae, Nymphaceae, Cruciferae, Aceraceae, in Phaseolus vulgaris and Lycopersicon esculentum (201–203). In roots, increase in total activity of SODs upon different heavy metals stress, such as Cu, Al, Pb, and Cu, was

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generally independent of plant species (soybean, lupin, Pisum sativum, Silene cucubalus) (196,204–207). Cadmium and copper caused an increase in total SOD activity in leaves of Pisum sativum and Phaseolus coccineus (197, 208, Sko´rzyn´ska-Polit et al., unpublished). Both metals and Zn decreased or had no influence on SOD activity in seedlings and leaves of Phaseolus vulgaris (62,198,209). Lower activity under Zn stress resulted from inhibition of Mn-SOD and/or FeSOD rather than from Cu/Zn-SOD (209). In Mn-tolerant genotype of Phaseolus vulgaris excess Mn caused increased leaf SOD activity, lower than in the susceptible one (210). Its excess also induced Mn-SOD activity in soybean and Pisum sativum (211–213). Palma et al. (208) showed an increase in peroxisomal MnSOD activity in Cu-tolerant Pisum sativum, suggesting that this dismutase may function in the molecular mechanisms of plant tolerance to Cu. Copper excess caused increase in Cu/Zn-SOD activity in seedlings of Triticum vulgaris, especially in the thylakoid-bound type of the enzyme (214). 5.2.2

Catalases

Despite the heavy metals applied (Cd, Cu, Zn, Pb, or Al—at higher concentrations), catalase (CAT) activity always decreased in roots of Phaseolus vulgaris, Lycopersicon esculentum, soybean, and lupin, although in roots of Allium cepa the enzyme activity was enhanced under Hg treatment (196,215–218). Somashekaraiah et al. (62) observed a decrease in CAT activity after 4 or 6 days of Cd action during germination of seedlings of Phaseolus vulgaris; however, in the seedlings of Phaseolus aureus (219) or in young plants of Phaseolus coccineus (Sko´rzyn´ska-Polit et al., unpublished) treated with Cd, an increase in CAT activity was reported. An increase in CAT activity was also measured in peroxisomes of Cu-treated Pisum sativum (208) and in leaves of Nicotiana plumbaginfolim under Fe excess (222). Short-term stimulation of CAT activity was also measured in leaves of Cu-treated Phaseolus vulgaris (217). Toxic concentrations of Mn and Zn increased CAT activity in celluler extracts from rice plants and in leaves of Pisum sativum (98,213). 5.2.3

Peroxidases

Peroxidases (POXs) have a much higher affinity to H2O2 than CATs (223). In higher plants, different isoenzymes of POX 3 can be distinguished. They participate in scavenging of H2O2, cell wall biosynthesis, regulation of IAA (indole3-acetic acid) degradation, synthesis of ethylene, and plant defense against pathogens. According to van Assche and Clijsters (48 and references therein), peroxidase induction is a general response of higher plants to uptake of heavy metals. The appearance of new isoperoxidases was observed when Oryza sativa was treated with Zn, Cu, and Hg. Zinc-dependent induction of two anionic POXs was observed in leaves but not in roots of Phaseolus vulgaris. However, Cu treatment of plants led to the reverse situation when Cd induced these isozymes

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in roots and leaves (48). An increase in nonspecific activity of POX was measured in lupin roots treated with Pb (196). Induction of guaiacol-dependent peroxidase was also observed in roots of Cu-treated Lycopersicon esculentum (217) and in soybean treated with Al (216). The appearance of two new anionic isozymes was observed in stems of Cd- or Zn-treated Phaseolus vulgaris where the induction of one of these iso-guaiacol POXs was Zn-specific (215). In seedlings of Phaseolus aureus (219), in leaves of Pisum sativum (197), and in both young and older plants of Phaseolus coccineus (Sko´rzyn´ska-Polit et al. unpublished), an increase in guaiacol-dependent POX was measured upon Cd stress. Chaoui et al. (215), in contrast to the results reported by van Assche and Clijsters (48), did not observe changes in POX activity in leaves of Cd-treated Phaseolus vulgaris. This discrepancy seems to be a result of the duration of Cd action and/or its various doses used in experiments. 5.2.3.1 Ascorbate Peroxidase Plants also have ascorbate-specific POX, which acts in cytosol and chloroplasts. This enzyme participates mainly in the ascorbate-glutathione cycle and plays a key role in scavenging H2O2 in chloroplasts. The activity of ascorbate peroxidase (POA) did not change in roots of Cd- and Zn-treated Phaseolus vulgaris (215), Cu-treated Lycopersicon esculentum (217), or decreased in lupin at higher concentrations of Pb (196). However, application of a lower Cu dose caused an increase in POA activity in roots of Phaseolus vulgaris (224). Dependence on metal concentration seems to be the case in Zn-treated Phaseolus vulgaris, where a lower concentration did not affect POA activity (215); however, at a higher metal content the activity increased (198). There have been several other reports of diverse responses of POA to heavy metals stress depending on plant species, specific organ, metal (Cd, Cu, Mn, Fe) and its concentration, plant sensitivity to metal, and experimental conditions (98,214,215,217,219–222). 5.3

Ascorbate/Glutathione Cycle

Ascorbic acid (AA) can be the electron donor to reduce H2O2 to water in reactions catalyzed by POA and, as mentioned earlier, participates in the ascorbateglutathione cycle, in addition to such enzymes as glutathione reductase (GR), dehydroascorbate reductase (DHAR), and monodehydroascorbate reductase (MDHAR). The influence of heavy metals on this cycle has been very poorly explored. In roots of Phaseolus vulgaris treated with excess Cu total ascorbic ⫹ dehydroascorbic acid (AA ⫹ DHA) and reduced ⫹ oxidized glutathione (GSH ⫹ GSSG) increased along with increase in all enzymes activity (MDHAR, DHAR, GR) (232). In leaves of sunflower treated with Cu, Cd, or Fe, decreased DHAR and GR activities and GSH content were observed (221). In Fe-treated Nicotiana plumbaginifolia a decrease in AA and GSH and an increase in DHA and GSSG was measured with no total increase in AA ⫹ DHA and GSH ⫹ GSSG (222).

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Ascorbate and GSH are themselves very good nonenzymatic scavengers. They react easily with free radicals (ROS, free radicals of proteins, lipids, and other compounds), and as a result of such reactions, the damage that occurred in different cell compounds of the cell may be repaired. Total pool of AA in tolerant type of Phaseolus vulgaris was maintained on a higher level under excess Mn stress than in the sensitive one (210). Yamaguchi et al. (225) showed a protective effect of endogenous GSH on Al toxicity in suspension-cultured tobacco cells. They suggested that GSH protects cells from oxidative membrane damage both by direct consumption of GSH and its oxidation. In all cases described in this chapter, symptoms of oxidative stress were observed in plants exposed to heavy metals stress. Their responses depended on plant species, age, tolerance or sensitivity to metals, time of exposure to stress, and concentration of heavy metal. The general condition of the photosynthetic apparatus seems to be related to ROS formation and, in consequence, competence and efficiency of the antioxidant system. 6

RESPONSES OF PLANTS TO HEAVY METALS: AVOIDANCE AND TOLERANCE MECHANISMS

Plants developed two types of strategies in response to heavy metals in the environment, i.e., stress avoidance and stress tolerance. Stress avoidance relies on reducing or preventing metal uptake by plants or its quick exclusion from protoplast by the excretion or sequestration mechanisms. Stress tolerance includes the mechanisms that deal with toxic metal ions present inside cells. 6.1

Avoidance

One of the strategies of stress avoidance is metal excretion from protoplast. An active excretion of metals (Fe, Cu, Zn) from leaf cells can occur via multicellular salt glands (Armeria maritima), hydathodes (Minuartia verna), or ectodesmata (Silene vulgaris) (226). Metabolism-dependent exclusion of Al from the root meristem of Al-tolerant Triticum aestivum was also shown (227). Restricted ion uptake can also result from plasma membrane lipids and sterols alterations, which is often described in bacteria, cyanobacteria, and algae. In higher plants this phenomenon does not play an important role, although it was shown in Cu-tolerant Silene cucubalus (66). A mechanical barrier against Al ions influx to protoplast is callose (1,3-β-glucan), deposited in the outer cortical cell layers of root tips. Its presence was noted among other factors in Glycine max, Zea mays, and Triticum vulgare (228). The cell wall is a mechanical and a chemical barrier against ion uptake as well as a compartment for deposition of ions that penetrated into the protoplast. Depending on the plant species and the kind of metal, it can retain from a small amount to more than 90% of the total metal taken up. Metals can be

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electrostatically bound to carboxylic functions of pectins and to charged groups of wall proteins, or even more complex chemical bonds can occur. Moreover, amorphous metal precipitates as phosphate, carbonate, and silicate deposited between cellulose micelles have been reported. The affinity of metal ions to polygalacturonic acid decreases in the order Pb ⬎ Cr ⬎ Cu ⬎ Ca ⬎ Zn (238). A similar reduction of uptake of Al and elements chemically similar to Al is caused out by mucilage, consisting mainly of polysaccharides and polyuronic acids (pectins) deposited on the root surface of, for example, Vigna unguiculata (230). Diminution of toxic ions uptake by plants is associated with rhizosphere modification by pH changes induced by plants or redox barrier in the plasma membrane of root cells. The redox barrier is formed as a result of oxidizing activity of plant roots and associated microorganisms. A number of wetland plants (Oryza sativa L., Spartina alterniflora, Phragmites communis, Typha latifolia) form so-called iron plaque on their roots by oxidizing Fe(II) to less soluble and less toxic Fe(III). A similar mechanism is present for Mn(II) immobilization in the rhizosphere (231). Mycorrhizae fungi can also be a mechanical and chemical barrier for heavy metal penetrating into plant roots (232). Plants also excrete metal-chelating ligands to the environment. Excretion of malic and citric acids to the rhizosphere by Helianthus annuus L., Triticum aestivum L., or polypeptides by Triticum aestivum L. in the presence of Al was observed (233–235). In Helianthus annuus L., excretion of malate and citrate was found in response to Zn (233). Heavy metals can be also deposited in idioblasts (226), hairs (236), and similar structures by which the heavy metal load can be withdrawn from plant metabolism. The phenomenon of seasonal intensive metal translocation to leaves and other plant parts and subsequent exclusion of metal from the organism by shedding these organs is also known. It occurs, for example, in Zn excluders Anthyllis vulneraria L. and Biscutella laevigate L., in Cu excluder Becium homblei (229,237), and in Ni-tolerant Indigofera setiflora (238). 6.2

Tolerance

The essence of intracellular mechanisms of tolerance is free heavy metal ions detoxification by their chelation in cytosol or storage in the vacuole. In most plant cells the vacuole comprises more than 80–90% of the cell volume and acts as a central storage compartment for ions. Some amino acids may also chelate metals in cytosol. The major ligands for Cu are asparagine (Asn) and histidine (His), and for Ni histidine, glutamine (Gln), and proline (Pro). Apart from their function in detoxification attention is paid to their role in metal transport in the xylem and phloem. Stability of these complexes can prevent precipitation of transported metals or their adsorption to cell walls (230,240). Organic acids are very effective chelators for toxic metal ions in cytosol and vacuole. The most important role in Zn tolerance is attributed to malic acid. According to the zincmalate shuttle hypothesis of Mathys (241), this organic acid is a cytosolic ligand

Angiosperms

193

for Zn, participating in the metal transport to the vacuole. In the vacuole, Znmalate complex is dissociated and malate is retransported into cytosol, whereas vacuolar Zn is bound to ligands more potent at low pH, such as citrate or oxalate, and also to phytate, anthocyanidines, or mustard oil glucosides, or is stored as Zn crystals. Zn-tolerance is also associated with enhanced production of citrate in the grass Deschampsia caespitosa (242). Both malate and citrate as well as malonate play a role in Ni detoxification (229,239). Organic acids are not a common means for Cu, Zn, and Pb detoxification and storage. An intensive synthesis and accumulation of high amounts of organic acids is not specific for the kind of metal or the plant species, and does not always correspond to tolerance to a given metal. For example, Zn-tolerant Agrostis capillaries and Silene vulgaris exhibit a high level of malic acid but are only merely Ni-tolerant. On the other hand, Ni-tolerant Alyssum bertolonii, rich in malate, is not Zn-tolerant (229). Another form of metal storage in the vacuole, especially of Zn but also of Cd, is chelation by phytin or phytate, a mixed salt of myoinositol hexaphosphoric acid or phytic acid. Metal complexes with phytin are found in seeds, cotyledons, leaves, and roots as dispersed in the protein matrix or aggregates, called globoids, and in immature cells as sheet-like deposits composed mainly of Cd and S (229,243). Studies on the role of proteins in heavy metal binding in plants were initiated per analogiam to such complexes, called metallothioneins (MTs) in animals. The first report about the presence of MTs in plants concerned Cu-MT in roots of Agrostis gigantea (244). Application of procedures appropriate for isolation and purification of such anionic complexes revealed that their structure and properties differ considerably from those of animal metallothioneins. That is why metal-protein complexes isolated from different plant species were no longer called MT but rather MT-like (245). In 1985, Grill et al. characterized Cd-binding peptides in Rauwolfia serpentina and named them phytochelatins (PCs) (246). PCs were designed to MTs class III and are defined as atypical nontranslationally synthesized metal thiolate polypeptides (247). They occur commonly in the whole plant kingdom (248) with the structure based on γ-Glu-Cys units repeated 2–11 times. The presence and kind of carboxy terminal amino acid were the criterion of distinguishing five families of γ-Glu-Cys peptides (240): Phytochelatins-(γ-Glu-Cys)n-Gly, found in all plant species studied Iso-PC (βAla)-(γ-Glu-Cys)n-βAla, found in some species of Fabales (250) Iso-PC (Ser)-(γ-Glu-Cys)n-Ser, found in Poaceae (251) Iso-PC (Glu)-(γ-Glu-Cys)n-Glu, found in Cd-treated maize roots (252) desGly-PC-(γ-Glu-Cys)n, found in species of Poaceae (251,252), Silene vulgaris and Lycopersicon peruvianum (253) Phytochelatins [(γ-Glu-Cys)n-Gly], which are glutathione derivatives, are the most widespread family responding at the earliest to increased intracellular

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metal concentration. That is why they are considered to play the main role in metal detoxification. Other families of γ-Glu-Cys peptides, although sometimes accumulated in greater abundance than PCs, play only an auxiliary though undoubtedly important role in the process of metal detoxification (251,254,255). The γ-Glu-Cys peptides are synthesized enzymatically by the action of a specific γ-Glu-Cys dipeptidyltranspeptidase trivially named phytochelatin synthase. This enzyme is activated by the presence of free metal ions, especially Cd(II), Ag(I), Bi(III), Pb(II), Zn(II), Cu(II), Hg(II), and Au(I). No enzyme activation was detected with Al(III), Ca(II), Fe(II), Mg(II), Mn(II), Na(I), or K(I) (256). A model for metal-dependent phytochelatin synthase function was recently proposed (257). Conversion of GSH into cadystins, structural analogues of plant phytochelatins, by the action of carboxypeptidase Y in vitro was also reported (258), but the capacity of metal ions binding by these peptides was not studied. Different functions for PCs in plant cells were proposed. PCs appear to be part of the homeostatic system that regulates the availability of Zn and Cu ions for apo forms of metal-requiring enzymes (259). They can also participate in assimilatory sulfate reduction as sulfo group acceptors from adenosine-5′phosphosulfate sulfotransferase (260). The most important function of PCs, however, seems to be their role in heavy metals detoxification and tolerance. Toxic metal ions present in cytosol induce rapid synthesis of PCs, binding metals in nontoxic complexes. It protects metal-sensitive groups such as -SH or histidyl groups of catalytic or structural proteins. Complexing by PCs is a mechanism especially important in Cd and Cu detoxification (261), but Pb and Zn, although they induce the formation of PCs, are not capable of forming stable complexes with them (253). First of all, efficient transport of metal-PC complexes to the vacuole plays an important role in plant tolerance to heavy metals. A model for PCs function in Cd transport to the vacuole was proposed by Vo¨geli-Lange and Wagner (262) and further developed by others (236,263). According to this, phytochelatins synthesized in cytosol bind Cd in so-called low molecular weight (LMW) complexes that are moved across the tonoplast by an ATP-binding cassette-type transporter. Inside the vacuole, more Cd, apo-PCs, and acid-labile sulfide are added to LMW complexes to form more stable sulfide-rich chelates called high molecular weight (HMW) complexes. Recently, more attention has been focused on searching and investigating MTs in plants. The intensity of those studies diminished after PCs characterization in plants (246). Although 64 genes predicting MT-like proteins from a variety of plants were already found and transcripts of some of them were detected in all plant organs, only one protein has been isolated and characterized so far (243 and references therein). It is Ec protein isolated from mature wheat embryos and belonging to MTs class II (i.e., peptides in which the position of the Cys residues is different from the archetypal mammalian MT and among each other) (247). Its presumable function is Zn(II) ions homeostasis. Whether this protein and other

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FIG. 4 Mechanisms of heavy metal tolerance and avoidance in plants.

MT-II gene products participate in heavy metals detoxification in plants remains to be examined. However, studies on the expression of plant MT genes in transgenic bacteria, cyanobacteria, yeast, and some plant species suggest that MT gene products can affect an increase in metal accumulation (especially Cd and Cu) in these organisms and enhance their metal tolerance (252 and references therein). The mechanisms of heavy metal tolerance and avoidance in plants are summarized in the Figure 4. 6.3

Biodiversity in Adaptation Mechanisms

Plant reaction to a given metal always involves many tolerance mechanisms and reveals that the predominant mechanism depends on plant species, growth phase

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or even plant organ or tissue, and the kind of metal, time of its action, concentration, and many other exo- and endogenous factors. However, some general tendencies for metals may be observed. Zinc is considered to be detoxified mainly by organic acids effectively transporting it from cytosol into the vacuole where it is stored. For Cd detoxification phytochelatins appear to be the most important, although some data suggest that the highest accumulation of this metal occurs in cell wall (264). The cell wall is the main compartment of Pb and sometimes Cu immobilization. An important role in Cu detoxification is also attributed to amino acids and metal-binding proteins. Such function for Ni is attributed to organic and amino acids, mainly histidine. The Poaceae family appears to be one of the most metal tolerant families in the plant kingdom. Apart from this, some grass species (e.g., Holcus lanatus, Agrostis capillaries, Festuca rubra, Deschampsia caespitosa) can very quickly adapt to metalliferous soils (265). Many tolerant ecotypes are also found in Caryophyllaceae, Brassicaceae (see Chapter 8), and Asteraceae. These families are especially widespread on metal-enriched soils of temperate zone, whereas in the tropics a much greater range of families is represented, including Fabaceae and Lamiaceae. However, it is interesting that some common and large plant families, such as Ranunculaceae, Leguminosae, Rosaceae, and Apiaceae, have not apparently evolved metal tolerance and only rarely occur naturally in metalenriched soils (265,266). The pattern of metal uptake, transport, and accumulation differs in each kind of metal or in various plant species and could be related to different mechanisms of metal tolerance at the whole-plant level. Three types of plant-soil relationships have been identified (267): Excluder strategy. Metal concentration in shoots is maintained at a constant low level until a certain critical soil concentration above which unrestricted metal transport to the shoots is observed. The shoot/root concentration ratio is much less than 1. Indicator strategy. Metal uptake and transport to shoots are regulated or passive uptake occurs so that the internal concentration reflects the external level. The shoot/soil concentration ratio is near 1. Accumulator strategy. Metal is actively concentrated within plant over the full range of soil concentrations. Analysis of plant organs suggests a general tendency for accumulators to translocate most metal taken up from roots to shoots, and the shoot/root concentration ratio is greater than 1. The term ‘‘hyperaccumulator’’ was first used by Brooks et al. (268) for plants accumulating more than 1000 µg/g (0.1%) Ni in their shoot dry mass when growing in the natural habitats. However, the threshold of hyperaccumulation

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varies considerably for different metals and amounts over 10,000 mg Zn and Mn (⬎1%); 1000 mg Ni, Cu, Co, Cr, and Pb (0.1%); 100 mg Cd (0.01%); and 1 mg Au (0.0001%) per 1 kg of plant shoot dry mass (238). Metal concentrations in some hyperaccumulators can be as high as 0.02% for Cd to 5% for Ni, Mn, and Zn (269). Unquestionably, the most impressive accumulator is an endemic tree Sebertia acuminata from New Caledonia in which 25.74% Ni was found in dried latex (239). It causes blue coloration of the tree, which is why its local name is ‘‘seve bleue.’’ Nowadays, more than 400 hyperaccumulators are known, including about 300 accumulators for Ni, 26 for Co, 24 for Cu, 19 for Se, 16 for Zn, 11 for Mn, 4–5 for Pb, 1 for Tl, and 1 for Cd (270). The extensive lists of hyperaccumulators were given by Baker and Brooks (238), Brooks (271), and Baker and Walker (237). In the temperate zone most hyperaccumulators belong to Brassicaceae (see Chapter 8); in the tropics to Euphorbiaceae. Hyperaccumulators can thrive in extremely hostile edaphic environments that would kill many other species. That is why they have been the subject of several investigations involving, among other things, their utility in detection of environmental pollution (see below), mineral prospecting (272), phytoarchaeology (using plants to search and investigate sites of ancient human activity) (273), phytoremediation (using plants to environment decontamination) (269), and phytomining (using plants to explore low-grade metal ores or highly mineralized soils) (270). 7

BIOMONITORING

Plants have been more and more widely used in the detection of heavy metals impact on the air, soil, and water caused by both natural processes and human activity. Assessing the environmental quality using plants as well as other living organisms is less expensive and often more efficient than using physical and chemical analyses. 7.1

Metallophytes

Due to developing different tolerance mechanisms (see above), some plants have adapted to the growth on metalliferous sites like mineralized soils: serpentine (rich in Ni, Cr, Mn, Mg, Co) or calamine (rich in Zn and Cd) soils, mineral deposits and outcrops, ancient and contemporary mining and smelting areas. Such plants are called metallophytes (274–276). Facultative metallophytes can grow both on metalliferous and nonmetalliferous soils. Examples of such plants are Anthoxanthum odoratum, Agrostis canina, A. capillaries, A. stolonifera, Deschampsia caespitosa, D. flexuosa, Festuca rubra, F. ovina, Holcus lanatus, and Silene vulgaris (274). Obligatory (or strict) metallophytes are strictly confined to metal-enriched soils, e.g., Haumaniastrum robertii, H. katangense, Viola

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calaminaria. That is probably due to their sensitivity to fungal attacks and inability to withstand interspecific competition on ‘‘normal soils’’ (276). 7.2

Indicators

Plants growing on heavy metals–enriched soils differ from those from surrounding sites. The vegetation structure and diversity is reduced. Usually there are no trees or the trees are stunted. Many xeromorphic features, dwarfed growth forms, decoloration of leaves (chlorosis, anthocyanous purple coloration), as well as other morphological changes can be observed in these plants. Endemic or disjunct species and edaphic ecotypes are of frequent occurrence (274,277,278). Many species are hyperaccumulators. That is why such plants can be used in biological monitoring of heavy metals pollution and also in mineral prospecting and phytoarchaelogy. In general, indicators can be divided into two groups—geobotanical and biogeochemical indicators. Biogeochemical indicators are used to detect mineralization by means of chemical analysis of the metal content in their tissues (277). Bioaccumulation studies seem most powerful in large-scale surveys when the expected date and localization of the pollution are unknown. They are the only possibility for detecting a chemical if its concentration in the environment is below the sensitivity level of the equipment (279) or is too low to cause characteristic symptoms of acute metal toxicity in plants (280). Geobotanical indicators are species that achieve their greatest abundance, occur exclusively or preferentially, or have a distinctive appearance (see above) or phenological pattern (delayed leaf flush, earlier flowering, premature senescence) on heavy metals– contaminated soils (277). A number of geobotanical and biogeochemical indicators have been reported so far (272, 276, 281, and references therein). Classic examples of geobotanical indicators are ‘‘copper flowers’’ of Shaba province, India—Becium homblei and Haumaniastrum katangense. Both grow only on soils rich in Cu (⬎ 0.01%) and used to be successfully used for prospecting for Cu deposits (275,276,282). Other ‘‘copper flowers,’’ or soil Cu indicators, include Haumaniastrum robertii, Bulbostylis pseudoperennis, Arthraxon quartinianus, Aeollanthus subacaulis var. linearis from South Central Africa (276,282), Polycorpaea spirostylis from Australia (283), Elsholtzia haichowensis from China (273). Silene cucubalus and Minuartia verna have colonized soils enriched by a variety of heavy metals, including Zn, Pb, Cd, and Ni in Europe (281). Plants used as geobotanical indicators represent all three strategies of plant-soil relationships, i.e., excluder, indicator, and accumulator, whereas biogeochemical indicators belong mainly to accumulators and indicators. The usefulness of vegetation in predicting the concentration of the elements in the soil decreases in the order herbs ⬎ shrubs ⬎ trees (272). It can probably be due to different radius of the root systems and distance for metal translocation between roots and leaves of herbs and trees. Plant organs most easily collected

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and analyzed for heavy metals content are leaves and also twigs or shoots, whereas roots, bark, and wood are difficult to collect. The pattern of metal accumulation in leaves and twigs varies depending on the metal. Aery and Tiagi (272) reported that leaves in the case of Zn and stems in the case of Pb, Cd, and Cu accumulated more of these metals than the others. If a plant occurring on metalliferous soils can be shown to be a hyperaccumulator of some element, although it is not confined to such substrates, the biogeochemical method of prospecting soil metals may be applicable. On the other hand, if a hyperaccumulator plant is also endemic to metalliferous soil, either absolutely or regionally, the geobotanical method of prospecting can be used. Biomonitoring of environmental pollution with heavy metals–sensitive plants can be also used. Morphological, physiological, and biochemical changes in these plants can give some information about poor environmental condition. Such a phenomenon can be easily observed in water ecosystems using, says, Lemna minor (284). 7.3

Biomonitoring in Experiments

Metal biomonitoring in the environment can be carried out not only studying individuals or communities in situ or samples taken from a natural population. Biological material may be also used in experiments (279). Some test plants of known physiology and reaction to heavy metals can be grown on polluted soil, both in the natural environment and under control conditions, to assess its toxicity. For instance, Phaseolus vulgaris L. cv. Limburgse vroege was successfully used to estimate the phytotoxicity of such soils. On the basis of its morphological (shoot length, leaf area, root weight) and physiological changes (the capacity of enzymes: peroxidase, POX; glutamate dehydrogenase, GDH; isocitrate dehydrogenase, ICDH; malic enzyme, ME), four classes of soil phytotoxicity were distinguished (280). Another test plant, commonly used to indicate Pb and other metals, e.g., Zn, Cd, Cu, Cr, Co, Hg is Lolium multiflorum var. italicum. Plants grown under standard conditions in unpolluted environment were exposed for 14 days to the air of the studied area and then are analyzed for concentrations of the individual elements (279). Solidago canadensis is used as an indicator of lead. Further terrestrial plants proposed for use are Anthriscus cerefolium, Medicago sativa, Raphanus sativus, Solanum lycopersicum, Volerinella locusta and Zea mays (279). In laboratory studies, the most widely used parameter to measure heavy metals toxicity/tolerance is root growth. This is because of the simplicity of this test and because this organ usually responds most rapidly to the metal and the response is the consequence of a direct effect of the metal on the root itself. Different variants of ‘‘rooting techniques’’ have been proposed, like IT (index of tolerance—the percentage of root elongation in metal solution compared with control solution), EC [the lowest effect concentration—the metal concentration at which root elongation is inhibited in 50% (EC50) or completely

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(EC100)] (229,274). Other parameters often used in metal tolerance tests include shoot elongation, water content, chlorophyll content, enzymes level and capacity, and many others (241,254,280,283). Adaptation to growth on heavy metal-enriched sites has some costs. It is accounted for by the fact that tolerant individuals, colonizing metal mine waste soils, have reduced fitness when grown on uncontaminated soils. It indicates the selection against metal tolerance on unpolluted soils. In general, tolerant mine plants are smaller and have lower biomass production and lower flowering in comparison with their nontolerant counterparts. Two hypotheses can explain how a cost of metal tolerance could be manifested. The ‘‘trade-off ’’ hypothesis suggests that the energy or resources used in the tolerance mechanism are diverted from other processes, e.g., growth and biomass production (285). In the ‘‘metal requirement’’ hypothesis tolerant plants are suggested to have an enhanced metal requirement, since tolerant individuals display a stimulatory response in growth and enzyme activity to essential metals (Cu, Zn, Ni), to which they are tolerant. On uncontaminated soil they are less efficient in the uptake, distribution, or utilization of these metals and their deficiency may occur. This argument forms the basis in the metal requirement hypothesis for explanation of the lower fitness of tolerant individuals when grown in normal soil (286). However, slow growth and other features of plants growing on metalliferous soils must not necessarily be a symptom of costs of tolerance to heavy metals. They may also result from adaptation to other stressful conditions in such environments, such as nutrient deficiency or drought (237,274). The survey presented here does not deal with all possible mechanisms of different heavy metals detoxification or tolerance in plants but rather with those reported most often. Among the tolerance mechanisms discussed in this chapter, changes in the plant metabolism to avoid metal toxicity inside the cell were not involved. These changes include activation of alternative metabolic pathways, increased enzyme synthesis, and changes in enzyme structures (274). Many taxon-specific responses to each heavy metal discussed have not been mentioned, but a great diversity in plant response to heavy metals has been demonstrated. ACKNOWLEDGMENT The financial assistance of the Polish Committee for Scientific Research (Grant 6P04C.064.15) is gratefully acknowledged. REFERENCES 1. LHP Jones, SC Jarvis. The fate of heavy metals. In: DJ Greenland, MHB Hayes, eds. The Chemistry of Soil Processes. New York: John Wiley and Sons, 1981, pp. 593–620.

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279. 280.

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Siedlecka et al. of the distribution of cadmium in roots of Zea mays L. J Plant Physiol 115:19– 28, 1984. MR Macnair, AJM Baker. Metal-tolerant plants: an evolutionary perspective. In: ME Farago, ed. Plants and the Chemical Elements. Biochemistry, Uptake, Tolerance and Toxicity. Weinheim: VCH, 1994, pp. 67–85. L Wu. Colonization and establishment of plants in contaminated environments. In: AJ Shaw, ed. Heavy Metal Tolerance in Plants: Evolutionary Aspects. Boca Raton: CRC Press, 1990, pp. 269–284. AJM Baker. Accumulators and excluders—strategies in the response of plants to heavy metals. J Plant Nutr 3:643–654, 1981. RR Brooks, J Lee, RD Reeves, T Jaffre´. Detection of nickeliferous rocks by analysis of herbarium specimens of indicator plants. J Geochem Explor 7:49–77, 1977. S Cunningham, DW Ow. Promises and prospects of phytoremediation. Plant Physiol 110:715–719, 1996. RR Brooks, MF Chambers, LJ Nicks, BH Robinson. Phytomining. Trends Plant Sci 3:359–362, 1998. RR Brooks. Plants that hyperaccumulate heavy metals. In: ME Farago, ed. Plants and the Chemical Elements. Biochemistry, Uptake, Tolerance and Toxicity. Weinheim: VCH, 1994, pp. 87–105. NC Aery, YD Tiagi. Bioindicators and accumulators in geobotanical and biogeochemical prospecting of metals. Acta Biol Hung 37:67–78, 1986. RR Brooks. Phytoarchaeology. Endeavour 13:129–134, 1989. H Schat, JAC Verkleij. Biological interactions: the role for non-woody plants in phytorestoration: possibilities to exploit adaptative heavy metal tolerance. In: J Vangronsveld, SD Cunningham, eds. Metal-Contaminated Soils: In Situ Inactivation and Phytorestoration. Austin: Landes Bioscience, 1998, pp. 51–65. A Baker, R Brooks, R Reeves. Growing for gold . . . and copper . . . and zinc. New Scientist 117:44–48, 1988. RR Brooks, F Malaisse. Metal-enriched sites of south central Africa. In: AJ Shaw, ed. Heavy Metal Tolerance in Plants: Evolutionary Aspects. Boca Raton: CRC Press, 1990, pp. 53–73. DE Wickland. Vegetation of heavy metal-contaminated soils in North America. In: AJ Shaw, ed. Heavy Metal Tolerance in Plants: Evolutionary Aspects. Boca Raton: CRC Press, 1990, pp. 39–51. AR Kruckeberg, AL Kruckeberg. Endemic metallophytes: their taxonomic, genetic, and evolutionary attributes. In: AJ Shaw, ed. Heavy Metal Tolerance in Plants: Evolutionary Aspects. Boca Raton: CRC Press, 1990, pp. 301–312. M Kova´cs, J Podani. Bioindication: a short review on the use of plants as indicators of heavy metals. Acta Biol Hung 37:19–29, 1986. J Vangronsveld, H Clijsters. A biological test system for the evaluation of metal phytotoxicity and immobilization by additives in metal contaminated soils. In: E Merian, W Haerdi, eds. Metal Compounds in Environment and Life. Interrelation Between Chemistry and Biology. Northwood: Science and Technology Letters, 1992, pp. 117–125. WHO Ernst. Mine vegetation in Europe. In: AJ Shaw, ed. Heavy Metal Tolerance in Plants: Evolutionary Aspects. Boca Raton: CRC Press, 1990, pp. 21–37.

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282. RR Brooks, AJM Baker, F Malaisse. Copper flowers. Natl Geogr Res Explor 8: 338–351, 1992. 283. AJM Baker, RR Brooks, AJ Pease, F Malaisse. Studies on copper and cobalt tolerance in three closely related taxa within the genus Silene L. (Caryophyllaceae) from Zaire. Plant Soil 73:377–385, 1983. 284. BS Mohan, BB Hosetti. Potential phytotoxicity of lead and cadmium to Lemna minor grown in sewage stabilization ponds. Environ Pollut 98:233–238, 1997. 285. FA Harper, SE Smith, MR Macnair. Where is the cost in copper tolerance in Mimulus guttatus? Testing the trade-off hypothesis. Funct Ecol 11:764–774, 1997. 286. FA Harper, SE Smith, MR Macnair. Can an increased copper requirement in copper-tolerant Mimulus guttatus explain the cost of tolerance? New Phytol 136:455– 467, 1997.

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8 Brassicaceae Luigi Sanita` di Toppi and Maria Augusta Favali University of Parma, Parma, Italy

Roberto Gabbrielli University of Florence, Florence, Italy

Patrizia Gremigni University of Western Australia, Perth, Western Australia, Australia

1

THE BRASSICACEAE FAMILY

The Brassicaceae (⫽ Cruciferae) are a typical example of a particularly homogeneous plant family, which comprises about 170 species already recognized by Linnaeus and termed ‘‘Tetradynamia.’’ The species belonging to this family are linked by a number of distinctive features, including the fairly uniform floral structure: the calyx and the corolla consist of four sepals (K2 ⫹ 2) and four petals (C4), respectively, which are arranged as a cross (hence the name Cruciferae). The stamens, four long and two short ones, form the androecium (A2 ⫹ 4), while the gynoecium is formed by two carpels (G2) and a superior ovary. The fruit, which also has an uniform structure, is a capsule (siliqua or silicle, depending on the length/breadth ratio) divided into two locula by a spurious septum which is not present in all species. Although the basic fruit structure is fairly uniform,

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it is characterized by a wide range of modifications that are relevant for classification at any level. The modern classification of the Brassicaceae is based on the revision by Schulz (1936) (1), who greatly contributed to the knowledge of this family in his work ‘‘Cruciferae,’’ published one year after his death. Janchen (1942) (2) published a rather thorough revision of Schulz’s system, in particular he reduced the number of the family’s tribes from 19 to 15. Nowadays the Brassicaceae family comprises 13 tribes. The Brassicaceae have a wide distribution and are present on almost every continent, mainly in the extratropical regions of the boreal hemisphere. In particular, the Alysseae and the Brassiceae tribes occur in several habitats, such as sunny and open places, walls, rock clefts, barren hills, fields, meadows, ruderal environments, river banks, and so forth. A number of Brassicaceae species occur in habitats with high anthropogenic pressure, and some species are of ornamental and horticultural importance. 1.1

Alyssum L.

The genus Alyssum L. is widespread in many regions of Europe and Asia. In Turkey, particularly in Anatolia, it achieves a high degree of diversity expressed by the greatest number of species. For the genus Alyssum, Dudley (1964) (3) reported six sections: Meniocus (Desv.), Psilomena (Meyer), Alyssum, Gamosepalum (Hausskn.), Tetradenia (Spach.), and Odontarrhena (Meyer). Dudley also reported a putative A. americanum (probably corresponding to A. obovatum that occurs in Siberia and in Turkey), which is indicated to be the only indigenous American representative of this genus. The genus Alyssum represents an interesting field of study for many botanists (4,5). However, at present the fame of Alyssum is not derived only from its taxonomic background but also from the ability of some species to survive on ultramafic substrates and to accumulate very high Ni concentrations. Already in the 16th century, Andrea Cesalpino (1583) (6), a doctor and botanist from Arezzo (Italy), described a ‘‘Lunaria quarta alias Alysson’’ as growing only over the ‘‘black stones’’ of Montauto, an ofiolitic outcrop of the Upper Tiber Valley in Tuscany. Later, Pier Antonio Micheli mentioned this plant as A. perenne in a manuscript on the flora near Florence. It was finally identified and classified as A. bertolonii by Desveaux (1814) (7). Then Amidei in the third meeting of the Italian Scientists (Florence 1841) (8) stressed the close relationship between some plant species and serpentine soils in Tuscany. However, in that time and up to the first half of the 20th century, the interest in some Alyssum species was more general. Many botanists were attracted by the peculiarity of the serpentine vegetation and its morphological alterations (i.e., dwarfism, glaucousness, etc.), mainly

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attributed to the extreme edaphic conditions to which these plants were adapted (9,10). 1.2

Brassica juncea

Brassica species have been cultivated for a long time and nowadays a number of varieties are widely used in various continents (11). Thus, as a result of this extensive spreading of mustard species throughout the world, there is a considerable variability for both morphological and biochemical traits (12). In particular, Brassica juncea (L.) Coss. and Czern. (Indian mustard) (Fig. 1) is an important crop species mainly in eastern Europe, India, Pakistan, China, and Japan. B. juncea has a great variability in the leaf morphology, and, consequently, it includes several cultivars. Some varieties are important as vegetables and for production of dressing oils. Most authors consider B. juncea to be of Asiatic origin, with centers of main biodiversity in China and perhaps in India. Morinaga (1934) (13) postulated different interrelation schemes among Brassica species and suggested that B. juncea has originated as a natural amphiploid hybrid (aploid chromosome number equal to 18) between B. nigra (L.) Kock (aploid chromosome number equal to 8) and other species with aploid chromosome number equal to 10, such as B. campestris L., B. rapa L., B. chinensis L., B. pekinensis Rupr., and B. japonica Sieb. Artificial synthesis of ‘‘pseudo-Juncea’’ forms, very close to the natural species, has been obtained by Frandsen (1943) (14) and by Ramanujam and Srinivaschar (1943) (15). 1.3

Thlaspi L.

The name Thlaspi was firstly used by Dioscorides Anarzabeus in his De medica materia (16,17) to describe a plant with specially shaped fruits (from the Greek thlao, to flatten, and aspis, shield). The genus Thlaspi includes about 75 herbaceous species with 30 annuals and 45 perennials (1,18). Thlaspi is primarily defined by fruit characters: ‘‘fruit an angustiseptate silicule, the valves keeled and usually winged; locules containing 2–6, rarely 1–10, seeds’’ (19), but the taxonomic classification of the genus is difficult and controversial (20). In his radical revision of the taxonomy of Thlaspi sensu lato (1,21), Meyer (17) analyzed the anatomy of the seed testa in taxa from Europe, Africa, and the Middle East. He split the genus into 12 segregate genera because the differences among them were too large to include all of them in one single, broad genus. More recently, Mummenhoff and colleagues used molecular techniques to study the controversial phylogenetic aspects of Thlaspi (19,20). The isoelectrofocusing (IEF) analysis of the enzyme Rubisco was applied to 14 species of Thlaspi (20). Preliminary results confirmed the existence of three major groups in the genus Thlaspi— Thlaspi sensu strictu (s. str.), Microthlaspi, and Noccaea (20)—consistent with

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FIG. 1 (a) Thlaspi caerulescens; (b) Brassica juncea

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three of the genera previously defined by Meyer (17). Thlaspi s. str. includes only annual species, whereas Microthlaspi comprises both annual and perennial species and supports Meyer’s results. Noccaea shows great heterogeneity and does not coincide with any previous systematic classification based on morphological and anatomical characters (20). Zunk et al. (19) investigated the genetic variation among 28 populations from metalliferous (subsp. caerulescens) and nonmetalliferous soils (subsp. calaminare) of the metallophyte T. caerulescens (Fig. 1). The isozyme analysis of aspartate aminotransferase, phosphoglucomutase, leucine aminopeptidase, and phosphoglucoisomerase indicated a very low degree of genetic variability in T. caerulescens and no correlation of genetic data with either the environmental factors such as Zn contamination or the geographic distribution of certain genotypes (19). The analysis of chloroplast DNA restriction site variation in 45 populations representing 20 Thlaspi species (22) and of sequences of the internal transcribed spacer regions of nuclear ribosomal DNA in 16 Thlaspi accessions (23) confirmed the presence of three major groups, thus supporting Meyer’s segregates: Thlaspi s. str., Microthlaspi, and Noccaea (including Raparia), with the greatest differences detected between Thlaspi s. str. and all other species (Table 1). Interestingly, the metal accumulating characteristics may be of taxonomic significance, since all the species belonging to Thlaspi s. str. are incapable of hyperaccumulating Ni or Zn, whereas several species of the lineages Noccaea and Raparia, which were originally part of Thlaspi [T. caerulescens J. et C. Presl., T. goesingense Hala´csy, T rotundifolium. (Wulfen) Kock subsp. cepaefolium (L) Greuter & Burdet, T. praecox, T. alpestre], are metal hyperaccumulators (19,24,25). However, the present chapter will refer to these Noccaea taxa as Thlaspi sensu lato (Table 1). The genus Thlaspi is distributed worldwide, primarily in the subartic and the north temperate zone of Eurasia, but also in eastern Africa, western North America and South America, northern Asia and Japan, whereas no species have been recorded in Australia (1,17,25,26). 2

HEAVY METAL ACCUMULATION AND DETOXIFICATION

There are about 300 Ni hyperaccumulator plant species, 26 Co hyperaccumulators, and 24 for Cu, 19 for Se, 16 for Zn, 11 for Mn, 2 for Tl, and 1 that hyperaccumulates Cd (27). Many Brassicaceae are known for their heavy metal accumulation and detoxification capability, in particular by species of the genera Alyssum, Brassica, and Thlaspi.

TABLE 1 Intrageneric Systems for Thlaspi s.l. Following Schultz (1), Clapham (26), and Meyer (17) Claphamb,c

T. sect. Nomisma DC. T. arvense T. sect. Chaunothlaspi O.E. Schultz T. alliaceum T. sect. Carpoceras DC. T. ceratocarpum

T. sect. Nomisma T. arvense T. sect. Nomisma

T. sect. Apterygium T. cepaefolium subsp. rotundifolium T. sect. Pterotropis DC. (incl. Neurotropis, DC.) T. montanum

T. sect. Apterygium T. cepaefolium subsp. rotundifolium T. sect. Thlaspi (sect. Pterotropis, Neurotropis) T. montanum T. caerulescens subsp. caerulescens subsp. calaminare T. macranthum T. alliaceum T. bulbosum

T. perfoliatum

T. perfoliatum

Thlaspi s. str. T. sect. Thlaspi T. arvense T. sect. Chaunothlaspi T. alliaceum T. sect. Carpoceras T. ceratocarpum Noccaea Moench N. sect. Noccaea N. rotundifolia L. Moench subsp. rotundifolia N. sect. Pterotropis N. montana (L.) F. K. Meyer N. caerulescens (J. & C. Presl) F. K. Meyer N. macrantha (Lipsky) F. K. Meyer Raparia F. K. Meyer R. bulbosa (Spruner) F. K. Meyer Microthlaspi F. K. Meyer M. perfoliatum (L.) F. K. Meyer M. natolicum (Boiss.) F. K. Meyer M. granatense (Boiss. & Reuter) F. K. Meyer

T. caerulescens and T. macranthum were not recognized by Schultz (1936). Thlaspi (Microthlaspi) granatense and T. (Microthlaspi) natolicum were not recognised by Schultz (1936) and Clapham (1964). c T. ceratocarpum does not occur in the area treated by Clapham (1964). d Meyer did not recognize subspecies with Noccaea caerulescens. Source: Modified from Ref. (23), courtesy of NRC Research Press, April 2000. b

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a

T. bulbosum

Meyerb,d

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Schultz a,b

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Alyssum

In 1948, while studying the mineral nutrition of the Impruneta ophiolitic outcrop vegetation (near Florence), as part of a geobotanical survey, two Florentine researchers, Minguzzi and Vergnano, discovered the unusual accumulation of Ni in A. bertolonii Desv. (28). Despite the discovery in Russia of a second ‘‘Ni plant,’’ A. murale (29), these results received little acknowledgment by the scientific community for a long time. Much later, after the discovery of other plants that are capable of accumulating high Ni concentrations (30–33) and especially thanks to the work of Brooks and co-workers and their strategy regarding the analysis of small herbarium samples (34), metal-accumulating plant species received further attention. At present, the term ‘‘hyperaccumulator,’’ coined to characterize the plants with Ni concentrations higher than 1000 µg g⫺1 in dry matter (34), is widely accepted. A wide investigation recognized that the genus Alyssum contains the greatest number about 48 of Ni hyperaccumulator species (35), 14 of which belong to the European flora (36). It is interesting to note that 13 of these European Alyssum and 22 Anatolian Ni hyperaccumulators (Table 2) were included in section Odontarrhena (35), demonstrating the combined high degree of endemism and resistance to high Ni concentrations of this group. On the other hand, it has been observed that the environmental pressure exerted on the adaptation capability of this genus may be due not only to high Ni concentrations but also to other factors causing the infertility of the serpentine soils. These include high levels of Cr, Co, and Mg, and a deficiency of nutrients like Ca, N, P, and K (42). Nevertheless, considering the similar effects observed in the same plants grown on ophiolitic substrates and in Ni culture solutions (43), and the several factors affecting the features of the serpentine soils, it is generally accepted that the presence of high Ni concentrations plays an important role in the evolution of a serpentine flora (28,44). A particular case is represented by Alyssum hyperaccumulators, in which a very high Ni tolerance has evolved. A. bertolonii showed a Ni tolerance higher than the one required by other nonaccumulator species that occur at the same ophiolitic outcrop (45). This suggests that metal hyperaccumulation must have a function besides the ability to survive on serpentine soils. Some hypotheses regarding the meaning of Ni accumulation have been proposed (46), but one of the most persuasive ecological explanations seems to be the Ni defensive role against herbivores or pathogens (47). This function, which might be similar in other hyperaccumulators, can be improved if the metal is localized in the outer layers of leaves and roots. As in other Ni accumulators, such as Hybantus floribundus (48), Senecio coronatus (49), and Thlaspi montanum var. siskiyouense (50), in A. bertolonii Ni has been evidenced in leaf epidermal cells as a red-stained Ni-dimethylglyoxime complex (51). Furthermore, microprobe analysis has shown that leaf hairs have the highest

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TABLE 2 Alyssum Hyperaccumulators

Species akamasicum Burtt alpestre L. anatolicum Hausskn argenteum All. bertolonii Desv subsp. Scutarium Nyar. callichroum Boiss. & Bu¨hse caricum T.R. Dudley & Huber-Morath cassium Boiss. chondrogynum Burtt cicilicum Bois & Balansa condensatum Boiss. &, Hausskn. constellatum Boiss crenulatum Boiss. cypricum Nyar. davisianum T.R. Dudley discolor T.R. Dudley & Huber-Morath dubertretii Gombault eriophyllum Boiss. & Hausskn euboeum Halacsy fallacinum Hausskn. floribundum Boiss. & Balansa

Ref.

Cyprus S. Europe (Alps) Anatolia Italy Italy (Tuscany) Italy Anatolia Anatolia Anatolia Cyprus Anatolia Iraq, Syria Anatolia, Iraq Anatolia, Syria Cyprus, Anatolia Anatolia Anatolia Anatolia Anatolia Euboea Crete Anatolia

9090 4480 8170 29400 13400 10200 10900 16500 20000 6300 13500 4090 18100 10400 23600 19600 11700 16500 11500 4550 3960 7700

35 36,37 35 35,37,38,39 28,36,37 39 35 35,39 35 35 35 35,39 35 35 35 35 35 35 35 36,39 36 35

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1 2 3 4 5 6 7 8 9 10 11 12 13 14 16 17 18 19 20 21 22 23

Location

Max. Ni in leaves (µg g⫺1 DW)

giosnanum Nyar heldreichii Hausskn. hubermorathi T.R. Dudley janchenii Nyar. lesbiacum (Candargy) Rech. f. malacitanum T.R. Dudley markgrafii O.E. Schulz. masmenaeum Boiss. murale Waldst. & Kit. obovatum (C.A. Mayer) oxicarpum Boiss. & Balansa peltarioides Boiss subsp. virgatiforme penjwinensis T.R. Dudley pinifolium Nyar. T.R. Dudley pintodasilvae T.R. Dudley pterocarpum T.R. Dudley Rech. f. robertianum Bernard ex Gren. & Godr. samariferum Boiss. & Hausskn singarense Boiss. & Hausskn. smolikanum Nyar. syriacum Nyar. tenium Halacsy (Tinos) trapeziforme Waldst. troodii Boiss. virgatum Nyar.

Anatolia Greece Anatolia Albania Lesvos Spain Albania Anatolia Balkans Russia Anatolia Anatolia Iraq Anatolia Portugal Anatolia Corsica Samar Iraq Grece Syria Tinos Anatolia Cyprus Anatolia

7390 12500 13500 9610 22400 10000 13700 24300 7080 4590 7290 7600 7860 21100 9000 22200 12500 18900 1280 6600 10200 3420 11900 17100 6230

35 36 35 35 35 34 35,39 35 29,35,36 35 35 35,39 35 35,39 36,40,41 35,39 35 35,39 35 36 35 36 35 35,39 35

Brassicaceae

24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48

Adapted courtesy of AB Academic Publishers, April 2000.

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Ni concentrations (52). Similarly, in A. lesbiacum, Ni was found in the epidermis and in leaf trichomes by a micro-PIXE technique (53). Probably the Ni accumulation capability is also closely connected to the degree of adaptation and restriction to ultramafic soils in order to avoid direct competition with the plants outside these sites (54). An increase in this very strong feature may have reduced the geographic spreading of Alyssum hyperaccumulators. This can be observed in the ultramafic soils of Turkey, where the distribution of the species of section Odontarrhena seems inversely related to its Ni concentration (35). The ability to concentrate metals in plant above-ground parts, reaching levels higher than the substrate metal concentrations, is a feature of metal accumulators and, to a greater extent, of the so-termed hyperaccumulators (55). Two important processes contribute to metal accumulation efficiency: a high root metal uptake and a high rate of long-distance transport. Among hyperaccumulators, Alyssum species show a general tendency to take up very high amounts of Ni. This metal is rapidly taken up from the serpentine soil and translocated to the leaves, where it is progressively stored. When Alyssum plants are grown in solutions with increasing Ni concentrations, the maximum accumulation level in the leaves is reached following a saturation trend (56,45). In A. bertolonii (28,57) and A. heldreichii (56), the Ni concentration is lower in the roots than in the leaves. However, the rate of Ni uptake in the roots is still higher in hyperaccumulator than in non-hyperaccumulator species. Ni concentration is also higher in the extractable fractions of ophiolitic soils than in normal soils (58). In the hyperaccumulators a great amount of Ni is immobilized in the cell walls of the root cortex due to its high cationic change capability. A decrease of pH can stimulate the release of this element into the apoplast. However, at present there is no evidence that the Ni fraction of the root cell wall contributes to the uptake into the cells. The relationship between plasma membrane transport and root uptake of metals has not been elucidated yet. Recently, evidence was found that Zn hyperaccumulation in T. caerulescens is associated with a plasma membrane transport mediated by specific types of selective channels (59). In Alyssum spp., besides the physical-chemical conditions favoring the release of Ni from the soil, a similar mechanism might support the preferential accumulation of this element rather than of other metals also present at high concentrations in serpentine soils. On the basis of a comparative study in A. troodii on the ability to accumulate Ni and Co, the same mechanism was suggested to be essential for the control of the competitive uptake of both the elements (60). Ni and Co have very similar chemical properties and both can be accumulated in leaves; however, Co accumulation in Alyssum plants is usually much lower than Ni accumulation. A highly selective uptake across the plasma membrane through an hypothetical ternary complex, like a protein-organic acid with tightly bound Ni, was proposed by Still and

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Williams (1980) (61). Great attention has been addressed to the chemical forms that ensure the nontoxicity of Ni ions. The mechanisms of Ni tolerance and accumulation can be interpreted as two aspects of the same process because the formation of stable nontoxic complexes is essential to Ni long-distance transport and storage in plant cells. The ability to bind Ni ions in organic acid complexes has been suggested to be a tolerance strategy for a number of Ni hyperaccumulator plants (62). In water extracts from A. bertolonii leaves, Ni was found to be present in the same gel filtration fraction containing mainly malic and malonic acids (63). Further evidence on the relationship between Ni and these organic acids was obtained comparing A. bertolonii grown in serpentine soils and in ordinary garden soils (64). Higher levels of both malic and malonic acids were present in the leaves of plants grown on serpentine soils, compared with plants from garden soil. In the Iberian hyperaccumulators A. pintodasilvae and A. malacitanum, Ni was separated as a polar complex associated with the same gel filtration fraction containing citric, malic, and malonic acid. By contrast, very small amounts of organic acids were present in A. serpillifolium, a nonaccumulator species closely related to the two Iberian hyperaccumulators (65). In all of these cases, the presence of both metal and organic acids in the same extract fractions has been asserted as accounting for the existence of a Ni complex, but this has not been supported by clear direct evidence. Nevertheless, since metals are mainly accumulated in the vacuoles (66), it is very likely that Ni in Alyssum is bound to organic acids. However, the formation of Ni complexes with other compounds, such as amino acids, could be favored by their higher stability constants, in sites where the organic acid concentration is much lower than in vacuoles (67). Kramer et al. (1996) (68) demonstrated that concentrations of Ni and histidine are highly correlated in the xilem sap of A. lesbiacum, A. murale, and A. bertolonii. The involvement of a histidine complex in the Ni accumulation process was directly supported by the X-ray absorption fine structure (EXAFS) analysis of xylem sap in A. lesbiacum. Further evidence of the involvement of this amino acid in Ni tolerance was given by the increase in biomass production and root elongation induced by exogenous histidine supplied to nontolerant A. montanum species cultured in solutions with toxic amounts of Ni. These findings suggest a mechanism of Ni accumulation in the genus Alyssum (Fig. 2), based on a nontoxic Ni-histidine complex for the transport of apoplastic Ni through the xylem sap to the leaves, where the metal is compartmentalized in the vacuoles by means of high concentrations of organic acids (62). It is also important to evaluate other factors and conditions that regulate the plant’s ability to accumulate Ni. In the hyperaccumulator species of Alyssum, different Ni contents have been observed (54). There is no easy way to evaluate if this characteristic is related to the natural genetic variability among species or

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FIG. 2 Nickel hyperaccumulation mechanisms in the ‘‘Ni pump’’ plant A. bertolonii.

if it is attributable to soil mineral composition. Nevertheless, different values of the maximum accumulation capacities were found in several Alyssum species grown in solutions with increasing Ni concentrations (56). In samples of A. bertolonii collected from an ophiolitic outcrop, the degree of Ni accumulation seemed to be related to the length of the growth period rather than to the Ni content of the soil (57). This relationship has been recently studied through mathematical modeling in A. bertolonii seedlings grown for 30 days in nutrient solutions (69). The suggested model of plant growth and Ni hyperaccumulation indicates that in A. bertolonii Ni uptake and tolerance are directly correlated only during the plant active growth phase. A. bertolonii accumulated up to 12,000 µg g⫺1 dry weight Ni in the leaves without showing the symptoms of growth inhibition. Interestingly, A. bertolonii plants grown in their natural habitat showed a similar upper limit of Ni accumulation. Another interesting aspect of Ni detoxification is represented by the regular shedding of the oldest leaves with the highest Ni content (57). This event represents a detoxifying process per se (70) and contributes to the establishment of the maximum metal accumulation level typical of each Alyssum species. 2.2

Brassica juncea

Dushenkov et al (1995) (71) reported the high capability of hydroponically cultivated roots of B. juncea to take up heavy metals up from nutrient solutions. For

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instance, roots of this species accumulated Cd and Pb, respectively, 134 and 563 times more than the initial concentrations in the solution. It was recently demonstrated that root cells of B. juncea mutants with increased Pb uptake have a higher amount of cell wall per gram of fresh weight than the normal root cells (72). According to the authors, this is an essential mechanism that allows the roots of this mutant to hyperaccumulate Pb. Cell wall desorption by EDTA further supported this hypothesis. B. juncea translocates Pb and other heavy metals from solutions (73) or soils (74) to shoots. It has been demonstrated that roots and shoots of B. juncea can accumulate various amounts of Pb according to the metal concentration supplied to the culture medium (75). Perhaps B. juncea shows a special capability to accumulate the highest levels of Pb in the shoot, as well as the ability to accumulate and detoxify significant amounts of Cd, Cr(VI), Cu, Ni, and Zn, compared with other heavy metal hyperaccumulators (73). In a comparative study between B. juncea and T. caerulescens, the former produced at least 20 times more biomass than the latter and showed a rapid growth rate under field conditions (76). However, T. caerulescens had a superior capability to accumulate all metals except Cd in the shoots (Table 3). The relative difference in metal root uptake between these two species was less dramatic than differences in shoot uptake. It is clear now that B. juncea first accumulates Pb in its roots and then translocates and concentrates the metal in the shoots (and hypocotyls) (77). In the presence of Cd ions, plants of B. juncea take up this metal by means of the root system and translocate part of it to the shoot, in particular to the leaves. Cd loading into the xylem sap of B. juncea displays saturation kinetics

TABLE 3 Shoot and Root Bioaccumulation Coefficientsa of Brassica juncea and Thlaspi caerulescens b Shoot Metal Cd (5)c Cu (1) Cr (0.4) Ni (1) Pb (5) Zn (3)

Brassica 175 159 80 587 3 49

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

16 32 8 115 1 31

Root Thlaspi

59 623 89 2739 29 770

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

12 265 15 383 23 320

Brassica 20574 55809 5486 11475 1432 1816

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

4295 9221 393 125 1409 1739

Thlaspi 4258 60716 8545 8425 7011 2990

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

168 21510 2677 4220 3616 1424

Bioaccumulation coefficient is the ratio of metal concentration in plant tissue (µg g⫺1 DW) to initial metal concentration in solution (mg l⫺1). b Hydroponically grown plants were exposed to metal solution for 8 days. c Initial concentration of metal in the solution (mg L⫺1). Source: Ref. 76. Courtesy of Nature Biotechnology, March 2000. a

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(78), suggesting that it is facilitated by membrane transporter systems (79). The movement of Cd in the xylem seems to be driven by the transpiration process (78). The accumulation of Cd in roots and leaves of B. juncea induces the synthesis of phytochelatins (78,80). The formation of Cd-phytochelatins complexes is the main system of Cd detoxification in leaves and roots of B. juncea. General mechanisms of Cd detoxification in higher plants have been reported by Sanita` di Toppi and Gabbrielli (81). EXAFS analysis in extracts from roots of B. juncea has demonstrated the Cd-S4 coordination. Differently, in the xylem sap Cd is complexed by oxygen and/or nitrogen: thus, Cd is more likely to be transported by organic acids than phytochelatins in the xylem sap (78). In the presence of Cd, glutathione synthetase (GS) is the rate-limiting enzyme for the synthesis of glutathione and phytochelatins in B. juncea. Therefore, overexpression of this enzyme might represent a powerful tool for the production of plants with the highest heavy metal accumulation capacity (82). Furthermore, the analysis of the expression of O-acetylserine(thiol)lyase (OAS-TL), γ-glutamylcysteine synthetase (γ-ECS), and GS in B. juncea has shown that the Cd-induced phytochelatin synthesis might be at least partially explained by the overexpression of glutathione biosynthesis genes (83). The Cd induction of a mitochondrial γ-ECS isoform in roots of B. juncea further supports this hypothesis (84). The response to Cd in roots and leaves of B. juncea is probably based on coordinated gene expression events tightly correlated to sulfur assimilation and glutathione synthesis (85). B. juncea is also an accumulator of Se (86). As briefly mentioned above with regard to Cd detoxification, high levels of S uptake and assimilation are also involved in Se metabolism in plants. As a matter of fact, Se-tolerant plants form acid-labile sulfide (S2⫺), and this has been demonstrated to be the mechanism for heavy metal detoxification also in B. juncea (80). Se toxicity might be counteracted by the incorporation of S-containing amino acids into proteins rather than their Se isologues (87,88). The incorporation of Se into proteins is probably the most important cause of Se toxicity in plants (89,90). In Se-nontolerant plant species, both selenocysteine and selenomethionine are formed and included in the elongating protein chain, where they replace the two S-containing amino acids cysteine and methionine. The presence of these ‘‘abnormal’’ amino acids in the proteins is responsible for Se toxicity in nontolerant plants (91). An increased Zn supply augmented the activities of guaiacol peroxidase, superoxide dismutase, catalase, and various components of the ascorbate-glutathione cycle in shoots of B. juncea seedlings. However, despite the prompt intervention of these detoxifying molecules, plants showed a significant increase in lipid peroxidation and tended not to survive (92). Interestingly, in Zn-stressed B. juncea plants, overexpression of glyoxalase I transcripts was also noticed. As reported by the authors, overexpression of glyoxalase I would diminish the level of methylglyoxal under heavy metal stress and allow the reformation of glutathione from its hemithioacetal (93).

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233

Thlaspi

Thlaspi heavy metal hyperaccumulating taxa have colonized specialized habitats, such as base metal deposits, undisturbed outcropping mineralizations, and heavy metal–polluted soils. Together with the typically heterogeneous vegetation of these metalliferous substrates, which reflects the wide variation of heavy metal concentrations over short distances, Thlaspi has probably developed to the present status within the last 150 years (94). However, T. goesingense is well represented on both mineralized and calcareous soils, whereas T. caerulescens and T. calaminare are represented only on metal-rich substrates and can be used as effective geobotanical indicator species in mineral exploration (95). The genus Thlaspi comprises Zn, Ni, and Pb hyperaccumulators (96) (Table 4). About 10 species have been identified as Zn hyperaccumulators (Table 4). The two closely related species T. calaminare (97,62) and T. caerulescens, which belong to the Zn-tolerant Galmei flora of western Germany and eastern Belgium (98–100), have been extensively studied (62). T. calaminare is one of the most important metallophytes of the Thlaspion calaminaris association confined to the calamine (Zn carbonate) deposits of western Europe. T. caerulescens has been the elective taxon for phytoremediation trials because of its efficiency in accumulating more than 4% dry weight Zn in the aerial parts (101,102). T. caerulescens, which now includes the variants originally grouped into T. alpestre (103), is also the most represented member of the so-called Minuartio-Thlaspietum alpestris association of the abandoned Pb mines of England, within the Pennines, Mendip Hills, and North Wales. Here the vegetation is affected not only by the high levels of soil Pb, Cd, and Zn, but also by soil instability, low organic matter status, and severe nitrogen and potassium deficiencies. Within the southern Pennines, the distribution of T. caerulescens strongly correlates with the location of Pb mines. Also in the mining areas of Mendip Hills T. caerulescens is the ubiquitous metal hyperaccumulator. T. caerulescens is also a well-known Cd hyperaccumulator species (more than 0.01% Cd in dried leaves) (62). Other Zn accumulators are T. rotundifolium subsp. cepaeifolium, widespread in a Pb/Zn mine of the Cave of Predil region (northern Italy) (104) and T. goesingense, which is also a Ni hyperaccumulator and is endemically restricted to a few serpentine outcrops in East Austria (105,106). Other five species of Thlaspi contain comparable amounts of Zn (more than 1% Zn in dried leaves), both in plants from Zn-rich substrates and from rocks with much lower levels of Zn (104). With a Pb content of 700 µg g⫺1 in dry shoots, T. rotundifolium subsp. cepaeifolium is the only relevant Pb accumulator species within the genus Thlaspi (Table 4) (104). This taxon has colonized the Raibl Mine (Pb/Zn mine) of the Cave of Predil region in northern Italy, with intermediate occurrence in other locations scattered around the confluence of the Freddo and Silizza rivers. The

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TABLE 4 Thlaspi Hyperaccumulators: Geographic Distribution and Maximum Metal Concentrations (% shoot dry weight) Species Zn Thlaspi alpestre L. brachypetalum Jordan bulbosum Spruner ex Boiss. caerulescens J. et C. Presl calaminare (Lej.) Lej. Et Court limosellifolium Reuter praecox Wulfen rotundifolium subsp. capaeifolium stenopterum Boiss. Et Reuter tatraense Zapal Pb T. rotundifolium subsp. cepaeifolium Ni Thlaspi alpestre L. subsp.virens (Jord.) Hook f. alpinum Crantz subsp. sylvium (Gaud.) Clapham bulbosum Spruner ex Boiss. cyprium Bornm. elegans Boiss. epirotum Hala´csy goesingense Hala´csy graecum Jordan japonicum Boiss. jaubertii Hedge montanum L. var. californicum var. montanum var. siskiyouense ochroleucum Boiss. Ex Heldr. oxyceras (Boiss.) Hedge rosulare Boiss. & Bal. rotundifolium (L.) Gaudin var. corymbosum (Gay) Gaudin sylvium (Gaud.) Clapham tymphaeum Haussknecht

Location

% Metal

Ref.

UK France Greece Germany Germany France Bulgaria Italy Spain Slovakia

2.50 2.00 1.05 2.73 3.96 1.10 2.10 2.10 1.60 2.70

98 98 98 98 98 98 98 105 98 98

Central Europe

0.16

105

Central Europe

0.40

105

Central Europe

0.41

105

Greece Cyprus Syria, Turkey Greece Austria Greece Japan Turkey

0.20 0.51 2.08 0.30 1.20 1.20 2.44 2.69

105 26 26 105 105 105 26 26

Western USA Western USA Western USA Greece Syria, Turkey Turkey Central Europe Central Europe Central Europe Greece, Albania

0.79 0.55 1.12 0.40 3.56 0.52 1.83 0.20 3.10 0.40

26 26 26 105 26 26 105 105 105 110

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mine tailings have been transported some distance from the river and accumulated at various sites downstream. The metallophytes are often found in these accumulations (62,104). T. rotundifolium subsp. cepaeifolium is a dominant member of the metal-lophyte plant community Thlaspietum cepaefolii associated with the Raibl Mine and described by Ernst (107,108). According to the author, the soil supporting this Thlaspi subspecies contains Pb (0.06–0.41%) and Zn (0.7– 14.6%), with small amounts of Cu (0–0.12%). Several Ni hyperaccumulator taxa of Thlaspi occur worldwide, particularly concentrated in the Mediterranean temperate regions and Turkey. T. goesingense Hala´csy is restricted to the Austrian Alps, where it is widespread on both metalliferous and nonmetalliferous soils (105,106). T. caerulescens is occasionally found also on serpentine soils in Scotland (109). At least three varieties of T. montanum that hyperaccumulate Ni (varieties californicum, montanum, and siskiyouense) were identified on the serpentine outcrops of North America (95). T. montanum californicum and siskiyouense are probably neoendemic forms, with variety montanum as a precursor (62,95). The genus Thlaspi includes both metal hyperaccumulators and nonaccumulators. The exceptional ability of the hyperaccumulator species to concentrate in their tissues a wide range of metals from both metalliferous soils and nutrient solutions has been extensively investigated (110,112,113). The physiological and genetic aspects of metal tolerance in Thlaspi are not fully understood, although they have been investigated in several Thlaspi species. 2.3.1

T. caerulescens

T. caerulescens is a very polymorphic species, with marked morphological differences among populations, sometimes even within the same geographic area (110). Wild populations of T. caerulescens from different locations of the British Isles were found to contain up to 21,000 µg g⫺1 Zn in dried shoots (110). When grown in nutrient solutions, plants accumulated up to 48,000 µg g⫺1 Zn and 1000–3000 µg g⫺1 Cd, without showing visible symptoms of metal toxicity (102). In a field trial with soil plots polluted with both Zn and Cd, T. caerulescens concentrated these two metals in the shoot tissue 10 times compared with the soil levels (111). With its extraordinary ability to hyperaccumulate Zn, Cd, and Ni, T. caerulescens (formerly T. alpestre L.) has become an ideal model for study of the mechanisms of heavy metal tolerance and hyperaccumulation in plants (112,113) and has a great potential in phytoremediation of polluted soils (104,114,115). Zn tolerance and hyperaccumulation are constitutive properties of T. caerulescens, although they are two physiologically and genetically independent traits (111,116–118). The mechanism used by T. caerulescens to mobilize soil Zn is not completely understood, and recent studies (119,120) show that rhizosphere acidification is not the main mechanism of Zn hyperaccumulation in this species. There is also little information on the mechanism of Zn transport and accumula-

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tion in the shoot, although the preferential accumulation of Zn in the epidermal cells of young and mature leaves has been conclusively demonstrated (121). Recently, Lasat and Kochian (112,113) used 65Zn2⫹ to quantify root fluxes, compartmentalization, and Zn translocation to the shoot in seedlings of the hyperaccumulator T. caerulescens and in the nonaccumulator T. arvense grown in nutrient solution. T. caerulescens accumulated significantly higher levels of Zn than T. arvense in the shoots, whereas the latter accumulated more Zn than T. caerulescens in the roots. However, both Zn accumulation in the roots and translocation to the shoot were faster in T. caerulescens than in T. arvense (113). The compartmentalization study at the root level resulted in similar amounts of Zn accumulated in the cell wall and cytoplasm of both species, whereas two- to fourfold more Zn was sequestered in the vacuole of T. arvense and made unavailable for translocation to the shoot. Furthermore, a higher Zn concentration was present in xylem sap of T. caerulescens and leaf sections of this species accumulated more Zn than those of T. arvense (113). These findings, together with the large deposits of Zn observed in the leaf vacuole of T. caerulescens (121,122), support the hypothesis that in the hyperaccumulator species a combination of stimulated transport systems, operating both at the leaf plasma membrane and the tonoplast level have different kinetic properties than in the nonaccumulator T. arvense (123). Molecular studies of Zn transport genes in T. caerulescens isolated and identified ZNT1 as one of the micronutrient transport genes, with high sequence homology with other Zn transport genes isolated from yeast (124) and from Arabidopsis (125). Zn hyperaccumulation in T. caerulescens relies on the ZNT1 gene, which encodes a high-affinity Zn transporter. This gene is constitutively expressed at a much higher level in T. caerulescens than in T. arvense, where its expression is stimulated by Zn deficiency. In fact, plant Zn status is shown to alter the normal regulation of Zn transporter genes in T. caerulescens. An important aspect of Zn hyperaccumulation and tolerance in T. caerulescens is also the production of low molecular weight compounds involved in Zn detoxification in the cell (cytoplasm and vacuole) and in the long-distance transport of Zn in the xylem vessels. Citrate was not shown to play an important role in Zn chelation, and malate reached constitutively high concentrations in the shoots of both the accumulator T. caerulescens and the nonaccumulator T. ochroleucum (118,126). More recently, direct measurements of the in vivo speciation of Zn in T. caerulescens using the noninvasive technique of X-ray absorption revealed that histidine is responsible for the transport of Zn within the cell, whereas organic acids (citrate and oxalate) chelate Zn during long-distance transport and storage (127). Another constitutive aspect of T. caerulescens is the high Zn requirement for maximum growth, compared with other species (118,120,128). This probably depends on the strong expression of the metal sequestration mechanism, which would subtract a large amount of intracellular Zn to normal physiological processes, even when the Zn supply is low (118).

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T. caerulescens, together with T. goesingense (129) was shown to have a great ability of accumulating the nonessential metal Cd (102,119,129–131). In a French population of T. caerulescens, Cd shoot concentration was exceptionally high (up to 10,000 µg g⫺1 dry weight) in hydroponically grown plants, and neither symptoms of Cd toxicity nor reduced plant biomass were observed (129). The current opinion is that nonessential metals can enter the plant cells only because of their chemical or physical similarities to plant nutrients, probably using Ca channels and broad-range metal transporters (132). The ability of T. caerulescens to accumulate a range of heavy metals, including Cd, cannot be explained any longer with the presence of a nonspecific uptake mechanism, side effect of an efficient mechanism of Zn hyperaccumulation, which can be adapted to different metals (110). Cd accumulation is a selective process, which does not share either common binding sites or similar compartmentalization mechanisms with Zn (119,130) and does not occur at the expense of a reduced Zn accumulation (102,129). The physiological and genetic basis of Cd tolerance in this species must be further clarified (102). Preliminary results suggest that in T. caerulescens Cd is stored mainly in the root apoplast and, to a lesser extent, in vacuoles, where it co-occurs with Ca and Fe (131). The highest efficiency in Cd hyperaccumulation may occur in populations characterized by more transporter sites at the root cell membrane (129). Differences in Cd tolerance observed among various populations correspond to differences in Cd hyperaccumulation ability, showing that Cd tolerance and accumulation are probably directly linked (129). Finally, the observed differences between Zn and Cd uptake in different varieties of T. caerulescens may be advantageous for improving phytoextraction of Cd through selection of plant populations, without losing efficiency in Zn phytoextraction (102,129). Some populations of T. caerulescens have an exceptional ability to take up other metals, such as Pb, Mo, Cr, Ag, Mn, Al, Co, Fe, and Cu, listed in increasing order of the biaccumulation coefficient (metal concentration in wholeplant dry matter/metal concentration in nutrient solutions) (111). Similarly to Zn, Cd, and Ni, Co and Mn absorbed by the root are easily translocated to the shoots, while the greater fraction of Cr, Cu, Al, Fe, and Pb is immobilized at the root level (111). 2.3.2

T. rotundifolium subsp. cepaeifolium

Previously termed T. cepaeifolium (133), T. rotundifolum is an effective hyperaccumulator of various metals (Cd, Ni, Pb, Zn) (104,106–108). Zn concentrations were found to be higher in leaves than in stems or roots (133). Interestingly, Wenzel and Jockwer (106) found that the amount of Cd, Pb, and Zn extracted from the soil by the best performing individuals of selected populations matched almost exactly the levels of phytoavailable metals in the soil (which is accessible also to nonaccumulating species). Based on these data, they suggested to change

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the lower limit of Cd concentration that qualifies plants as Cd hyperaccumulators from 100 (134) to 50 µg g⫺1 dry weight (106). 2.3.3

T. goesingense

Two different varieties of T. goesingense are known. One occurs on Austrian ultramafic (serpentine) outcrops and is capable of hyperaccumulating up to 12,500 µg g⫺1 Ni dry matter in the shoot (105,106). The other variety is well represented on calcareous soils and is also a Ni hyperaccumulator when grown on ultramafic soils (24). The ability to hyperaccumulate Ni in the shoots is thought to be a constitutive property of this species (111). Recently, Kra¨mer et al. (135) compared the rate of Ni transport and tolerance in the Ni hyperaccumulator T. goesingense and the related nonaccumulator T. arvense. They found that the hyperaccumulating phenotype observed in hydroponic cultures did not depend on different rates of Ni transport from root to shoot in the two varieties, but probably on a specific cellular mechanism of Ni tolerance in the leaves of the hyperaccumulator (135). On the other hand, the exudation of histidine and citrate in the rhizosphere, which plays an important role in Ni detoxification in Alyssum hyperaccumulators (135), is not essential for Ni hyperaccumulation in T. goesingense, whereas the exudation may be responsible for decreasing Ni uptake and toxicity in T. arvense and other nonaccumulator species (136). Furthermore, Pearsans et al. (137) demonstrated that, although genes involved in histidine biosynthesis are expressed at the root and shoot level, they are not regulated by Ni concentrations and time of Ni application. The concentration of histidine in root, shoot, and xylem sap after Ni exposure is similar in T. goesingense and T. arvense, and so it is when histidine is measured directly in plant tissues of both Ni accumulator and nonaccumulator species (137). The role of histidine in the detoxification and transport of Ni, as suggested in the Ni hyperaccumulator A. lesbiacum (138), was not demonstrated in T. goesingense and T. arvense. T. goesingense is very efficient in accumulating also Cd concentrations up to 830 µg g⫺1 in dried shoots (129), well above the concentration usually considered to qualify a Cd hyperaccumulator (100 µg g⫺1) (134). 2.3.4

T. montanum

The species T. montanum includes the Ni hyperaccumulating varieties montanum, californicum, and siskiyouense, which occur on serpentine soils of North America. Also T. montanum var. fendleri is reported to accumulate both Ni and Mn (139). T. montanum var. montanum is peculiar because it occurs both on serpentine and normal soils (140). Similarly to T. goesingense, in T. montanum var. montanum metal hyperaccumulation is a constitutive trait, which confers to the plant adaptive advantages on metalliferous soils (140). The nonserpentine populations are ‘‘latent hyperaccumulators’’: with their relatively high Ni concentrations in the leaves (3–345 µg g⫺1 Ni in dried shoots) on nonmetalliferous

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239

soils, they show a predisposition for metal hyperaccumulation and can easily hyperaccumulate metals on metalliferous soils (140). Since most Ni hyperaccumulators show also high levels of Zn in their shoots, Ni hyperaccumulation may be the consequence of an effective nutrient uptake mechanism on fertile soils, which can also be used with Ni when this species occurs on Ni-rich soils (140). One of the advantages of accumulating Ni in the leaves is an efficient defense against predators and pathogens, which can be killed by the toxic effect of Ni (141–143). There is clear evidence that the Ni hyperaccumulator T. montanum var. siskiyouense accumulates the highest Ni concentration at the leaf surface and particularly in the subsidiary cells surrounding guard cells and in other elongate epidermal cells, and this also supports the defense hypothesis (144). 2.3.5

T. ochroleucum

T. ochroleucum is a pioneer plant on some metalliferous soils (24) and is highly tolerant to such heavy metals as Ni and Zn; however, it is not a hyperaccumulator (111,120,145). In T. ochroleucum Zn is accumulated in the root and Zn tolerance is lower than in T. caerulescens due to a less efficient root-to-shoot transport of the metal (118). This ecotype very rarely occurs on Cu-rich soils, and tolerance to Ni and Zn does not confer to T. ochroleucum tolerance to Cu. Cu has adverse effects on germination, seedling survival, and seedling growth (146) and causes severe damage primarily at the root level (147). 2.3.6

T. sylvestre Jord. subsp. calaminare

T. sylvestre Jord. subsp. calaminare occurs on the Belgian calamine soils, accumulates Zn and requires high levels of Zn for germination and normal growth (20 and 50 mg L⫺1 in nutrient solution, respectively) (148). 2.3.7

T. arvense

T. arvense is a nonaccumulating species that has been widely used to compare the performance of metal hyperaccumulators and non-hyperaccumulators (113,149, 150). T. arvense is a widespread weed in Eurasia and North America and is a pioneer plant on abandoned lands. It is an aggressive competitor that causes severe losses in agricultural crops as well as a popular food plant, especially in Europe (16). With its succulent leaves, it is capable of surviving in extreme drought conditions, probably due to a typical CAM metabolism (151). 3

PHYTOREMEDIATION

For many years, the species that accumulate unusual amounts of heavy metals have been considered by several scientists as potential indicators of metalliferous soils and interesting models in which to study the capacity of plants to adapt to extreme environmental conditions. In the last decade, the increasing interest in

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phytoremediation, whereby certain plants are used to remediate contaminated sites, has also emphasized the great potential of metal hyperaccumulators (152). These plants could represent an appropriate and cost-effective tool for removing heavy metals from polluted sites. Furthermore, in the case of the termed ‘‘phytomining’’ practice, they can give an economic return through the recovery of the metals from low-grade ores (62). The efficiency of phytoremediation of heavy metal–contaminated soils requires that plants have (a) great ability to take up and translocate metals; (b) high efficiency in detoxifying the metal in plant tissues; (c) high production of above-ground biomass. Many authors used the total metal translocated (metal concentration in shoot biomass/dry weight) as a measure of the effectiveness of metal hyperaccumulating plants in extracting heavy metals from contaminated soils. Among Brassicaceae there are several species of potential high interest for phytoremediation. 3.1

Alyssum

The potential use of A. bertolonii in soil phytoremediation has been evaluated in terms of the conditions capable of improving its efficiency in taking up and accumulating Ni ions (153). In fact, reduced growth rate and biomass production, which are usually associated to Ni tolerance and accumulation (154), may limit the use of many metal-hyperaccumulating species to clean up contaminated soils. A. bertolonii plants grown on an ultramafic soil and treated with appropriate fertilizers (N, P, K), had an increased biomass productivity without changes in Ni concentration. By contrast, the addition of calcium carbonate, which reduced Ni and Mg availability, decreased the Ni content in A. bertolonii plants. Generally speaking, Alyssum hyperaccumulators have been considered useful tools, capable of competing economically with traditional practices of heavy metal decontamination (153,155). Alyssum species appear to be interesting tools also in remediating sites contaminated with other heavy metals, in particular Co, that can be usually found in ultramafic soils together with Ni (155). To optimize the use of Alyssum in metal recovery, appropriate conditions have been individuated: (a) lowering of the soil pH; (b) maintenance of low calcium levels or lowering of calcium levels in the soil by leaching; (c) use of ammonium-containing or ammonium-generating nitrogen fertilizers; (d) distribution of chelating agents to the soil (155). 3.2

Brassica juncea

About 1 g dry weight from roots of B. juncea, placed in 400 ml of water supplied with 300 µg ml⫺1 of Pb, reduced Pb levels in the solution to less than 1 µg ml⫺1 in a few hours. Simultaneously, the concentration of Pb in the roots exceeded 10% of the tissue dry weight (156). In other experiments, B. juncea plants were shown to accumulate Pb in the root system (75) or in stems and leaves (157).

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Kumar et al. (158) suggested that B. juncea could respond even further to increasing Pb levels. In particular, the Vardan genotype concentrated Pb in different plant parts, showing a better ability in Pb uptake than other genotypes. Consequently, the Vardan genotype seems to be the most suitable for Pb phytoextraction. However, for phytoremediation of Pb-contaminated soils, the chelateassisted phytoextraction strategy (79) appears to be in general more effective than the natural heavy metal phytoextractive ability of Pb hyperaccumulators. EDTA promptly chelates Pb ions in the soil, then the Pb-EDTA complex is taken up by B. juncea roots and translocated, following the transpiration stream, to the leaves, where it is accumulated at very high levels (159). The use of chelating agents (EDTA, EGTA, etc.) might contribute to a more effective phytoremediation of Pb-contaminated soils. Interestingly, Epstein et al. (160) observed significantly higher levels of Pb in B. juncea plants grown on a soil containing 4.8 mmol kg⫺1 Pb and treated with 1 mmol kg⫺1 EDTA than in plants from a soil containing 1.5 mmol kg⫺1 Pb and treated with 5 mmol kg⫺1 EDTA. The authors conclude that a higher ratio of Pb to EDTA in the soil solution will result in more Pb taken up by the plant. However, the addition of EDTA to metal-polluted soils has the undesirable effect of reducing the plant growth significantly. For instance, B. juncea plants grown in the presence of 5 mmol kg⫺1 EDTA in the soil produced a twofold greater biomass than plants treated with 10 mmol kg⫺1 EDTA. Furthermore, at the higher EDTA concentration, the leaves started wilting, had a transpiration rate diminished by 80%, and had a relevant abscission (160). Interestingly, several strains of Pseudomonas and Bacillus stimulated Cd accumulation in B. juncea seedlings from hydroponic solutions (76). Pilon-Smits et al. (161,162) overexpressed the E. coli γ-glutamylcysteine synthetase (γ-ECS) and glutathione synthetase (GS) enzymes in B. juncea. These transgenic plants concentrated more Cd than normal plants in their shoots. They might represent a strong tool for developing effective Cd phytoremediation strategies (82,83), possibly also toward Pb and Hg ions. Differently, scarce improvement of Cd accumulation in B. juncea with manipulated sulfate metabolism has been reported in experiments from Lee and Leustek (163). These authors conclude that ‘‘changes in gene expression may not always be indicative of a physiologically relevant adaptive response to an environmental condition,’’ in full agreement with the concepts discussed by Sanita` di Toppi and Gabbrielli (81) and Meharg (164). Brassica spp., which tend to accumulate S, have also high ability to take up Se, due to the chemical similarity between SO42⫺ and SeO42⫺. Se is usually present in the environment under the following forms: elemental Se, selenide (Se2⫺), selenite (Se4⫹), selenate (Se6⫹), and organic Se (selenocysteine and selenomethionine; see Sec. 2.2) (165). Furthermore, dimethylselenide and dimethyldiselenide are forms of volatile Se (166). The oxidized forms, SeO32⫺ and SeO42⫺,

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are soluble in water and consequently fairly available to plants. Possible variations in the metabolism of Se in different varieties of B. juncea should be taken into account when screening for a specific variety able to clean up Se-contaminated soils (167). In general, B. juncea leaves accumulate Se to the highest concentration when Se is supplied as SeO42⫺, followed by selenomethionine and SeO32⫺, whereas selenomethionine accumulation is favored in roots, followed by SeO32⫺ and SeO42⫺. In addition, Se volatilization is closely correlated to Se accumulation in roots (168). In general, Brassica spp. may improve their ability of accumulating Se if high levels of SO42⫺ are present in the soil solution. B. juncea seems to be ‘‘one of the best plant species for Se phytoremediation because it grows rapidly, produce a large biomass and efficiently volatilizes and accumulates Se’’ (169,170). Phytoremediation of Se-contaminated sites involves accumulation of Se in plant tissues (171), together with volatilization of Se both by plants (172) and by soil microrganisms. On average, Se can be accumulated in B. juncea tissues up to 1000 mg kg⫺1 dry weight (166). Interestingly, B. juncea cleaned up a SeO42⫺-contaminated soil in 3 years, removing up to 50% of Se within a depth of 75 cm (173). Recent studies have shown that SeO42⫺ accumulation in B. juncea root system may be rate limiting for Se assimilation and volatilization, which at the same time can be efficiently stimulated by rhizosphere bacteria (174,175). Since bacteria appear to increase Se accumulation in B. juncea only when this heavy metal is supplied in the chemical form of SeO42⫺ (not in the form of SeO32⫺ or selenomethionine), it is likely that different Se uptake mechanisms are expressed in B. juncea plants (176–179). Thus, B. juncea plants in association with some rhizosphere bacteria seem to be good ‘‘joined forces’’ in phytoremediation of SeO42⫺-contaminated sites. In phytoremediation of Cr(VI)-contaminated soils, the reduction in planta of the toxic Cr(VI) to the less toxic Cr(III) may be useful. Preliminary results obtained by means of X-ray absorbance spectroscopy (XAS) revealed that roots of B. juncea, as well as other mustard species (196), are capable of reducing Cr(VI) to Cr(III). Thus, B. juncea may be of relevant interest also in remediating Cr-polluted soils. Marked potential in U hyperaccumulation in the presence of citric acid has been shown in B. juncea, B. narinosa, and B. chinensis. U accumulation is a fast process: U can be found in the shoots within 24 h after the treatment with citric acid, and it reaches a maximum in about 3 days. B. juncea plants, when used for U phytoextraction, could be harvested just a few days after being supplied with citric acid. As reported by the authors, with this technique U accumulation in shoots of B. juncea raised by more than 1000-fold within a few days. They conclude that ‘‘applying this technique in the field will speed up the removal of U from contaminated soils and will provide a cost-effective soil decontamination strategy’’ (180).

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B. juncea may also be effective in phytoremediation of Zn-contaminated soils. Interestingly, Ebbs et al. (181,182) used barley and B. juncea simultaneously, since the former had the capability of hyperaccumulating very high levels of Zn and the latter, when Zn was supplied in association with EDTA, was able in turn to accumulate Zn not previously taken up by barley. Finally, Anderson et al. (27,183) reported hyperaccumulation of Au in B. juncea plants cultivated in Au-treated pots in the presence of ammonium thiocyanate. In these experimental trials, interesting preliminary results of possible Au phytoextraction were obtained. 3.3

Thlaspi

Thlaspi has attracted the attention of many scientists because several hyperaccumulating species within this genus are suitable for phytoremediation of soils polluted with high levels of Zn, Cd, Ni, and Pb. A preliminary hydroponic or pot screening needs to be developed to identify the best accessions within a species, in terms of metal accumulation and best potential performances in phytoremediation, to overcome the great variability seldom observed among populations of metal hyperaccumulators of the same species (102,181). The superiority of T. caerulescens toward other plants was demonstrated and explained with its extraordinary efficiency in sequestering metals in the shoot (101,114). Furthermore, this species offers the possibility of multiple metal removal due to its proven ability to accumulate high levels of Zn, Cd, and Ni. Initially, the time span required for complete or limited removal of metal contaminants from the soil was calculated assuming that subsequent crops would remove metals at the same rates as the first crop (111,114). Baker et al. (111) calculated that about 13 annual croppings were needed to reduce soil Zn content of a polluted soil from 444 to 300 µg g⫺1 (limit set by the Commission of the European Community, 1986). However, if the metal dissolution rate in the soil decreases below the potential extraction rate by the plant, due to possible changes in metal partitioning in the soil, the effectiveness of phytoremediation may decrease with time (106). In a different approach, Robinson et al. (184) calculated the total number of annual croppings of T. caerulescens required for removing half of the soil metal, based on the assumption that this was the maximum amount that could be extracted by the plant. Soils contaminated with less than 500 µg g⫺1 Zn or 20 µg g⫺1 Cd could be remediated to half these values in about 10 and 2 years, respectively (184). Certainly the duration of metal phytoextraction is still an overly long process in this case to be commercially attractive (111,185). Although T. caerulescens is very efficient in taking up high levels of Zn, Cd, and Ni from the soil as well as accumulating these metals in the shoots, the use of this species for commercial scale phytoremediation (115,152) is limited by its small size (about 15 cm high), slow growth rate, and rosette growth habit

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(185,186). The mechanical harvest of these plants is very difficult, the dry weight yield is too low (6-month-old plants of T. caerulescens yielded about 3–4 t ha⫺1 dry weight) (111,185). One possible solution to this shortcoming might be the transfer of Zn-hyperaccumulating properties from T. caerulescens into a high biomass-producing plant (101) as a realistic possibility of making phytoremediation a commercial technology. Other species of the genus have been shown to be very efficient in accumulating Zn from Zn polluted soils in a greater plant biomass (185) but perhaps not particularly resistant to the high levels of metal accumulated (187). On the other hand, the use of molecular techniques to transfer the desired genes for metal hyperaccumulation from T. caerulescens into larger plants is premature because, despite the fact that research has been done in this direction, genes responsible for plant metal tolerance and hyperaccumulation have not been fully identified in metal hyperaccumulators. Studies of the genetics of metal hyperaccumulation in T. caerulescens have shown that Zn tolerance and hyperaccumulation are separate traits (117) and generally metal tolerance is a polygenic trait in other species of plant accumulators (188). Another possibility, which has been successfully explored in two laboratories (186,187), is the production of somatic hybrids through fusion of protoplasts from parents bringing the desired traits. Sexual incompatibilities between parents with widely divergent backgrounds can be overcome, even though the somatic hybrids obtained might be sterile or not viable. But the few fertile hybrids that survive can be backcrossed to one parent, and genes can be transferred between different genomes via nonhomologous recombination (189). Most of the intertribal somatic hybrids that were produced between T. caerulescens and Brassica napus (186,187) had intermediate size. One of the two fertile hybrids obtained (hybrid 60/31) showed both the desired characters of vigorous growth and highly increased resistance to Pb, and the amount of Pb that this hybrid was able to extract from the soil was similar to that of both parents (187). Also, some Zn-tolerant hybrids with fast growth and erect growth habit were obtained and will be back-crossed with B. napus to get viable seed (186). Attempts to increase plant biomass production by growing T. caerulescens in a more fertile environment (190,191) were not always successful. Different varieties of T. caerulescens grown in pot and field experiments had different responses to soil fertilizer application: one population produced significantly higher plant biomass and extracted more Zn and Cd from polluted soils amended with N, whereas the other produced more biomass but had reduced Zn and Cd uptake when grown either on P or on N fertilized soils (191). In other field experiments with T. caerulescens (190), the plant proved to be adapted to growing in an ultramafic environment. Only the addition of N fertilizer increased plant biomass, but it slightly reduced Ni and Zn concentrations in the shoot. Furthermore, Ni and Zn levels in plant tissues were much lower than those found in wild plants, probably due to a too short growing period and to the high growth rates achieved

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under the experimental conditions. However, T. caerulescens showed a potential in phytoremediation of soil Ni (190). Furthermore, the application of metal complexing agents such as EDTA to contaminated soils to enhance metal phytoextraction (181,192), which has been successful in increasing metal uptake in B. juncea (74), might be used to increase metal uptake by T. caerulescens as well (193). T. rotundifolium subsp. cepaefolium is a well-known Pb hyperaccumulator, which accumulates up to 8200 µg g⫺1 Pb in the shoots (24). However, it is not suitable for Pb phytoextraction because of its very low growth rate. Recently, B. juncea has been found to have a more efficient translocation system that concentrates up to 34.5 mg g⫺1 Pb in the shoot (73). Progress in the area of soil phytoremediation is closely dependent on a better understanding of the physiological mechanisms involved in Zn transport and accumulation in root and shoot of hyperaccumulating plants. While several studies on metal long-distance transport and storage in the leaf of T. caerulescens have been published in the past few decades, only recently did Schwartz et al. (194) and Whiting et al. (195) highlight the importance of root development and function in T. caerulescens. In this species, roots showed a high affinity for Zn and exhibited two major types of morphologies, depending on the presence of Zn in the soil, with “cluster” roots in the presence of high Zn levels and finer and longer roots in noncontaminated soils (194). T. caerulescens root architecture seems to depend specifically on Zn form, concentration, and localization in the soil. Plant roots tend to explore a rather large soil volume and colonize preferentially soil ‘‘spots’’ with high Zn concentration to ensure that the plant’s high requirements for Zn are satisfied (194). Similar root proliferation toward localized soil Zn enrichments was observed in T. caerulescens but not in T. arvense by Whiting et al. (195). In the nonaccumulator T. arvense roots were shorter and smaller than in the hyperaccumulator T. caerulescens and tended to proliferate toward the noncontaminated soil. The discovery of these sophisticated Zn-scavenging mechanisms in T. caerulescens roots is essential for a better understanding of metal acquisition in plants. Future screenings for the selection of hyperaccumulating plants suitable for phytoremediation will have to take into account not only plants with high shoot biomass but also those with high root/shoot mass ratio and high specific root length (195). The trait “metal hyperaccumulator” may result from efficient root foraging for metals and preferential production of root biomass toward soil metal sources. ACKNOWLEDGMENTS Thanks to Drs. H. Lambers, M. Tomaselli and L. Pignotti for their helpful revision and advice. We are also grateful to Drs. C. Gonnelli and E. Torricelli and to Miss C. Padovani for their thorough assistance with the artwork.

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9 Aquatic Macrophytes M. N. V. Prasad University of Hyderabad, Hyderabad, India

Maria Greger Stockholm University, Stockholm, Sweden

Bruce N. Smith Brigham Young University, Provo, Utah

1

INTRODUCTION

Angiospermous families, such as Cyperaceae, Potamogetonaceae, Ranunculaceae, Typhaceae, Haloragaceae, Hydrocharitaceae, Najadaceae, Juncaceae, Pontederiaceae, Zosterophyllaceae, Lemnaceae, Typhaceae, etc., have aquatic and semiaquatic habitats. While a great diversity of freshwater aquatic plants carry out their complete life cycle floating on the surface (e.g., Eichhornia and Lemna), others live on the surface but retain an attachment to the sediments (Nymphaea). Many others are emergent with only the roots submerged (Ranunculus, Typha, Carex, etc.). A few species live completely submerged (Potamogeton, Najas, Ruppia, etc.) while retaining a connection to sediments from which minerals are absorbed (Fig. 1a,b). All aquatic angiosperms have terrestrial ancesters 259

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but, like dolphins and whales, have adapted to life in the water (1). While minerals can be absorbed from water by leaves, for many plants rooted in sediments, root uptake is paramount. Macrophytes take up metals from water, producing an internal concentration several fold greater than their surroundings. They play a crucial role in biogeochemical cycling of trace metals. The submerged plant thickets in polluted lakes are reported to accumulate trace metals to the tune of 103 –104 or reduce the water velocity there by accelerating sedimentation of the suspended fine particulate trace metals, which otherwise are toxic to the biota when present in the interstitial waters in available form (2,3).

2

METAL UPTAKE BY DIFFERENT SPECIES OF AQUATIC PLANTS

Heavy metals are enriched in all compartments of the environment (air, water, and soil) by anthropogenic activities (4). Finally, the aquatic systems act as receptacle for several of these metals. For example, the metal concentration in river waters increased several thousand fold by effluents from mining wastes. Background levels of metals in freshwater and seawater, as well as in sediments of marine and lacustrine origin, are depicted in Figs. 2 and 3 (5).

FIG. 1 Vertical zonation in a tropical wetland plant assemblage. (a) A generalized representation of zonation consisting of seven zones that occur at the land-water interface and the various life forms of plant species that occupy each zone, including the position of above-ground structures and roots in relation to the sediment surface and water table. (b) A typical zonation in tropical wetland macrophytes. Floating islands consisting of Eichhornia crassipes, Lemma, Wolfia, etc. species, euhydrophytic zone comprising landemergent plants species encompassing Typha latifolia, Nymphea, rooted submerged Vallisneria spiralis; and submerged free-floating, e.g., Ceratophyllum demersum. Aquatic macrophytes can be broadly divided into three categories: (i) Free-floating plants: Except roots, the whole body of the plant is above the water, e.g., Eichhornia crassipes (water hyacinth), Ludwigia sp. (water primrose), Pistia stratiotes (water lettuce), Lemma sp., Wolffia, and Spirodela, etc. (duckweeds), Ceratophyllum demersum (coontail), Salvinia sp. and Azolla (water ferns). (ii) Submerged (rooted) plants: These remain submerged in water, e.g., Egeria densa (brazilian elodea), Elodea canadensis (elodea), and Hydrilla, Vallisneria (tape grass). (iii) Emergent (rooted) plants: These are rooted to sediments but emergent above the water, e.g., Alternanthera philoxeroides (alligator weed), Typha latifolia (cattail), Phragmites communis (reed).

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FIG. 2 Background concentration of metals in seawater and freshwater (5).

Many environmental factors are known to modify the availability of metals in water to aquatic plants. Such factors are chemical speciation of the metal, pH, organic chelators, humic substances, particles and complexing agents, presence of other metals and anions, ionic strength, temperature, salinity, light intensity, and oxygen level, and thereby the redox potential (Eh) (Fig. 4) (4). In lakes, pH is important for speciation and thus also for the availability of metals to macrophytes. Most metal concentrations in water increase with decreasing pH, with the highest pH value of about 4. Especially in aquatic systems, the redox potential (Eh) is important. At low redox potential, metals become bound to sulfides in sediments and are thus immobilized. In water, salinity affects the availability of metals, since high salinity causes formation of metal-chloride complexes, which are hard for plants and other organisms to take up (5–6). An equilibrium is maintained in the interface between the metals in water and metals in the sediment (Figs. 5 and 6). Plants are able to take up metals from these media. Sediment are considered as sinks and metals are therefore accumulated in sediments in high concentrations. Table 1 details the terminology used for transfer of metals from sediments to interstitial waters and plants. Water and air function more or less as transport media for elements. It is known that

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FIG. 3 Background concentration of metals in marine and fluvial sediments (5).

metals have a shorter retention time in air and water compared with that of sediments. However, this depends on the element. For example, Pb has a very short retention time in water, whereas for Zn it is longer. Therefore, plant uptake of metal as well as the distribution within the plant tissues is influenced by the retention time. Furthermore, there is an equilibrium in the interface between the metals in the water and the metals in the sediment. This means that if some external factor influences the equilibrium, the uptake of metals by plants may also be affected (7–9). Such factors include resuspension, bioturbation, addition of organic matter, and changed salinity. It has been shown that salinity decreased the uptake of Cd, Cu, and Zn in submerged plants in hydroponic conditions (9). However, if the system also included sediment, salinity gave an opposite effect, increasing the Cd uptake in the submerged plant. The explanation was that in the first case, CdClx complexes were formed in the water, and these complexes are hard for the plant to take up. In the second case, addition of salt released Cd bound in the sediment, probably by exchange with Cd and Na on the colloid, thereby changing the equilibrium toward a higher Cd concentration in the water. Thus, uptake of Cd by the shoot increased (9).

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FIG. 4 Model of metal (Me) translocation in freshwater macrophyte-watersediment (9).

FIG. 5 Model of metal (Me) translocation in saline environment (9).

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FIG. 6 Various useful methods for monitoring heavy metals using aquatic macrophytes.

Different plant species take up different amounts of metals (9–19) (Tables 2–6). The concentration found in various plant materials varies from site to site due to environmental pollution situation at that site. The variation within the same species can therefore be large (Tables 2 and 5). There are also differences in uptake between various metals. Not only the uptake but also the toxicity of

TABLE 1 Transfer of Metal from Sediments to Interstitial Waters and Biota (33) Term Bioconcentration Bioconcentration factor

Bioaccumulation Biomagnification Biominification

Explanation Concentration in organisms greater than concentration in medium, i.e., concentration factor ⬎1 (Shoot concentration ⫻ shoot biomass ⫹ root concentration ⫻ root biomass)/[(shoot biomass ⫹ root biomass) ⫻ duration of growth in days] (28) Increase in concentration with age or size of organisms Increase in concentration in organisms at successively higher trophic levels Decrease in concentration in organisms at successively higher trophic level

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TABLE 2 Concentration Range of Lead and Copper in Different Macrophytes Collected from 11 Sites of Two English Lakes, Ullswater and Coniston Water (20) Cu (µg g⫺1 dry wt.)

Pb (µg g⫺1 dry wt.)

Macrophyte

Shoot

Root

Shoot

Root

Potamogeton crispus Potamogeton perfoliatus Potamogeton, other spp. Myriophyllum alterniflorum Elodea canadensis Littorella uniflora Isoetes lacustris

19–37 13–27 14–510 8–490 9–460 14–80 19–90

18–60 10–16 13–550 11–165 250–490 18–140 45–330

95–500 25–180 37–950 26–590 18–930 7–310 15–1200

140–6600 71–700 31–2400 29–1000 34–110 35–1500 40–5700

metals varies. Toxicity of different metals to Elodea canadensis and Myriophyllum spicatum is highest with Cu ⬎ Hg ⬎ As ⬎ Cd ⬎ Zn ⬎ Pb (8). Submerged plant leaves have a very thin cuticle. The leaves of submerged plants are therefore very good at metal uptake directly from the water. The foliar uptake of Cd by Potamogeton pectinatus was nearly 10 times higher than that of Pisum sativum (9). Welsh and Denny (20) showed that lead was accumulated in the shoots via absorption from the water. Macrophytes take up heavy metals via roots from the sediment and via shoots directly from the water. Therefore, the integrated amounts of bioavailable metals in water and sediment can be indicated by using macrophytes (Fig. 7). Plants can also evolve ecotypes pretty soon and thereby encounter unfavorable conditions. Plants are also stationary, long-lived, and accumulate metals; therefore, they are suitable in monitoring of polluted sites. Metal concentration in

TABLE 3 Concentration Factors for Some Elements in Macrophytes Element Macrophyte Elodea Myriophyllum Potamogeton Najas Nuphar Phragmites

M

Zn

Fe

Co

Cu

6190 2670 1760 1080 530 1770

1950 1600 740 560 570 350

8770 6280 4064 3513 1268 3388

18.0 19.9 8.0 3.8 1.7 2.5

189 181 98 45 64 52

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TABLE 4 Leaf Tissue Mn and Pb Concentrations (µg g⫺1 dry wt.) of Representative Roadside Plant Species from Moderately High Traffic Volume Roadsidesa Plant species Aquatic Ranunculus aquatilis Potamogeton pectinatus Ruppia maritima Lemna minor Scirpus acutes Chara spp. Herbaceous Asclepias speciosa Kochia scoparia Cardaria draba Convolvulus arvensis Lactuca serriola Medicago sativa Graminaceous Distichlis spicata Agropyron cristatum Agropyron repens

n

Manganese

Lead

3 5 3 3 3 3

13,680 3,414 1,755 569 393 489

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

1,153 240 339 81 6 23

8.0 15.6 56.3 14.5 8.0 37.0

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

0.1 3.3 8.2 1.1 2.1 2.8

3 13 3 7 6 5

117.5 106.2 85.0 97.7 43.8 44.2

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

9.6 14.3 26.7 21.7 7.7 3.1

5.0 5.2 14.7 1.6 5.3 10.2

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

1.4 1.8 2.8 0.8 3.1 4.3

3 26 7

104.0 ⫾ 11.3 42.7 ⫾ 3.4 46.1 ⫾ 7.4

3.0 ⫾ 2.1 7.1 ⫾ 1.4 12.6 ⫾ 2.2

Each value represents the mean ⫾ SE of n ⱖ 3 plant samples. Source: Data from Ref. 91.

a

plants must be related to the time of the year since the metal concentrations vary in season (21). Aquatic plants also release metals through their leaves. Plants that are used as bioindicators must retain the metals in their plant body. Ceratophyllum demersum, Myriophyllum spicatum, Potamogeton pectinatus, P. perfoliatus, and Zannichellia palustris have been proposed as bioindicators (22–28). 3

REDISTRIBUTION DUE TO UPTAKE BY SEDIMENT

Bottom sediments have been polluted with heavy metals by industrial outlets over decades. Due to the pollutants, but probably mainly due to reduced light intensity caused by increased density of particles in the water column, the ability of plants to grow on these bottoms will be low. The metals have been bound to organic particles and sedimentated to the bottom. In the sediment, the redox potential is low and therefore most metals are firmly bound to sulfides. When the industrial outlet has been shut down, new unpolluted sediment has been deposited on top of the contaminated one and plants are able to colonize

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TABLE 5 Concentration Range of Heavy Metals in Different Macrophytes Collected from 47 Sites of Shallow Costal Areas, Stockholm, Sweden Cd Species Potamogeton perfoliatus Elodea canadensis Ranunculus baudotii Myriophyllum spicatum Potamogeton pectinatus

Cu

Pb

Zn

Ni

Cr

S

R

S

R

S

R

S

R

S

R

S

R

0.7–2.5 1.3–1.7 2.0–2.8 0.7–3.3 0.5–2.9

0.9–4.0 1.8–2.6 2.6–4.7 0.4–7.3 0.6–2.7

7.3–58 21–22 9.9–37 4.5–32 3.8–26

3.3–10 18–26 14–75 6.7–34 2.8–33

74–520 84–93 88–377 65–386 43–171

20–719 66–106 112–124 35–185 17–106

0.3–51 6–23.6 9.8–17 0.3–57 0.3–54

0.3–73 7.5–109 25–32 0.7–148 0.3–115

3.7–28 34–46 6.5–12 4.9–25 5.5–17

2.3–30 14–77 8.0–14 5.8–50 3.9–34

0.05–5.7 1.6–3.2 0.05–0.9 0.05–4.7 0.3–3.1

0.05–4.6 0.8–6.3 1.0–1.7 0.1–6.4 0.05–4.4

S, shoot; R, root. Source: Data from Ref. 19.

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(mean ⫾ SE, n ⬎ 3) in Utah, USAa Category Areas Provo River Utah Lake Bear River Sevier River Clear Lake Fish Springs Plants Ranunculus aquatilis Potamogeton crispus P. filiformis P. pectinatus Ruppia maritima Najas marina Carex spp. Scripus maritimus Distichlis spicata Chara spp.

Cadmium

Iron

Lead

Manganese

2 3 0.7 0.5 0 0.8

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

0.2b 0.2b 0.2a 0.5a 0a 0.2a

1649 1748 489 948 93 127

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

277b 125b 79ab 392ab 13a 17a

17 65 7 1 10 5

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

4b 11c 1ab 0a 1ab 1a

3660 445 177 70 51 18

⫾ ⫾ ⫾ ⫾ ⫾ ⫾

682b 143a 71a 13a 11a 2a

3 3 2 2 0.6 2 0.3 1.3 0.3 1

⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾

0ab 0.2ab 0a 0.2a 0.2a 0.7a 0.3a 0.2a 0.3a 0.4a

756 1035 4686 818 437 89 72 43 80 157

⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾

321ab 226b 367b 110ab 173a 25a 9a 19a 12a 36a

8 21 12 12 10 8 1 9 1 8

⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾

3a 11a 2a 3a 3a 3a 0.2a 1a 0.4a 2a

13,680 3741 3552 1458 243 23 40 305 27 10

⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾ ⫾

1632b 2014a 271a 330a 137c 23c 17c 66c 11c 3c

Aquatic Macrophytes

TABLE 6 Aquatic Plant Tissue Metal Concentrations (ppm dry wt.) by Species and Wetland

a Listed in approximate descending order of human impact. Means within columns not sharing a letter are significantly different, P ⬎ 0.05. Source: Data from Ref. 94.

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FIG. 7 Factors regulating metal bioavailability in sediments.

the bottom area. Colonization of macrophytes on shallow bottoms brings about increased redox potential in the rhizosphere caused by the photosynthetic oxygen which has been translocated down to the roots by the lacunar system (29–31). The increased redox potential increases the bioavailability of heavy metals to the plant roots, thus facilitating the uptake of heavy metals by the roots. Part of the metals up taken will then be translocated to the leaves and thereafter transferred to grazing animals. Thus, macrophytes can increase the metal circulation in the aquatic environment. Several macrophytes not only serve as indicators of metal pollution but also redistribute the metals in the aquatic ecosystem (32–39).

4

FACTORS INFLUENCING METAL UPTAKE

The uptake of metals, both by roots and by leaves, increases with increasing metal concentration in the external medium. However, the uptake is not linear in correlation to the concentration increase. This is due to the fact that the metals are bound in the tissue, causing saturation that is governed by the rate at which the metal is conducted away. The uptake efficiency (or accumulation factor) is highest at low external concentrations. This is shown for Cd in solution culture (22). Greger also demonstrated increased uptake from one and the same metal concentration with increasing root absorption area (root mass). [External factors, such as temperature and light, not only influence growth but also affect metal uptake (Fig. 4).] In aquatic macrophytes, the metal removal is accomplished through uptake by roots, chemical precipitation, and ion exchange with or absorption of settled clay and organic compounds (40–42). Metal uptake is enhanced due to the pres-

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ence of metal-binding ligands such as thiols or synthesis of metal-chelating peptides/proteins, i.e., phytochelatins (43–51) or surfactants such as linear alkylbenzenesulfonate (52), or combinations of metals in wastewaters may exert synergistic or antagonistic influence on metal uptake (53). Aquatic sediments have a low oxygen content, and since metals are more soluble and readily taken up in anoxic and low pH environments, aquatic macrophytes typically have much higher metal contents than nearby terrestrial plants, even when the total metal content of the soils is equivalent. Risk assessment of toxic trace metals in aquatic biota and use of inbuilt water-renovating strategies on ecologically acceptable principles has gained considerable significance in the field of environmental biotechnology and biotechnological methods of pollution control (54–57). Aquatic macrophytes such as water hyacinth (Eichhornia crassipes) and several duckweeds have attracted the attention of scientists for their ability to accumulate trace metals. Using E. crassipes, heavy metal uptake from metalpolluted water, metal specificities on introduction and binding affinities of heavy metal–binding complexes in root, purification and characterization of the metalbinding complex, and sorption of heavy metals from metal-containing solutions have been extensively investigated (58). Metals found in nature in more than one valance state are more readily taken up by plants in the reduced form. Often the metal is reoxidized within the plant tissues. People studying mineral nutrition in plants have usually distinguished mobile and immobile elements. Heavy metals, or even lighter metals in excess, are often toxic to plants. Even at sublethal concentrations, physiological tests, such as changes in pigment composition, photosynthesis, and respiration, can reflect stress and predict future plant damage (59–61). 5

WATER-SEDIMENT INTERACTIONS, METAL AVAILABILITY, AND PARTITIONING

Sediments (particularly lacustrine) act as an important reservoir for metals of anthropogenic source. The fluxes of metals to sediments may be ‘‘gravitational,’’ i.e., settling of particulate material with its associated metals, or diffusive, i.e., in cases where dissolved metal concentrations in the water column exceed those in the sediments interstitial waters, notably in acid lakes. Thus, in the upper strata of lake sediments, higher levels of metals would be available than those in the overlying water column (32–39). In aquatic ecosystems, the sediments and water interface serve as the habitat for a diverse community of whatever aquatic macrophytes are prominent. Metals present in surficial sediments in particulate form may exist either as constituent elements present in the essentially insoluble products of physical weathering form, i.e., lattice-bound (metal in detrital or residual minerals), or in a variety

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of secondary forms [adsorbed on surfaces (iron/manganese oxides, clays, humic flocs)], associated with organic form matter, sulfides, etc. The secondary forms are more reactive and are more likely to be bioavailable. The cardinal factors affecting the metal bioavailability to macrophytes in aquatic ecosystems are temperature, salinity, pH, Eh, and metal-metal interactions (Fig. 7). Aquatic macrophytes have paramount significance in the monitoring of metals in aquatic ecosystems (62–65). In wetland ecosystems, a wide variety of processes ranging from physicochemical to biological operate that can provide a suitable situation for removal of metals (Fig. 7). For example, in the case of acidic metal-rich mine drainage, the principal processes include oxidation of dissolved metal ions and subsequent precipitation of metal hydroxides, bacterial reduction of sulfate and precipitation of metal sulfides, coprecipitation of metals with iron hydroxides, adsorption of metals onto precipitated hydroxides, adsorption of metals onto organic or clay substrates, and, finally, metal uptake by growing macrophytes. 6

CONSTRUCTED WETLANDS FOR REMOVAL OF METALS WITH MACROPHYTES

Constructed wetlands with reed beds and floating-plant systems have been common for the treatment of various types of wastewaters for many years. This strategy is currently gaining importance globally and expanding to address contaminated/polluted soils and water bodies (Table 5). Bioconcentration of trace metals by aquatic macrophytes is of special concern to human welfare and for environmental protection and conservation (66–68). Natural wetland ecosystems are inherently complex. Hence, for the purpose of treatment of metal-contaminated waters it is advantageous to construct separate tanks within the treatment system, with each tank designated to perform a particular function maximally (occasionally more than one would be beneficial). The design of wetlands constructed for the treatment of metal-contaminated waters represents an attempt to identify and optimize the key processes that promote the removal of a specific targeted metal. Alternatively, this includes suppression of potentially interfering and competing processes. Treatment of wastewaters/natural waters containing a single metal, such as iron, can be achieved using a constructed wetland designated to optimize only one of the possible factors. For example, removal of iron involves precipitation of iron hydroxide in an aerobic environment. In contrast, if the water contains a mixture of metals, e.g., iron and zinc, in high concentrations, the constructed wetland may have to adapt different strategies, such as application of aerobic and anaerobic processes. An aerobic environment promotes the precipitation of aluminum and iron hydroxides and coprecipitation of arsenic. An anaerobic situa-

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tion promotes the reduction of sulfates and the consequent precipitation of sulfides primarily for copper, cadmium, and zinc. The precipitation of hydroxides is regulated by pH and the availability of oxygen, which can be ensured as follows: 1. Construct shallow wetlands with a maximal water depth of about 3 m. 2. Minimize use of organic detritus because it demands oxygen for decomposition. It is preferable to use large inorganic substrate. 3. Design the landscaping into ridges and gullies to ensure continual mixing of the water within the system so as to prevent stratification of water into oxygen-rich and oxygen-depleted zones. 4. Incorporate cascades (Fig. 8) at the point of influent to promote oxygenation of air. 5. Utilizing reed beds comprising Phragmites australis (common reed), Typha latifolia (cattail), etc., which have the ability to transfer oxygen to the root zone (Fig. 9). Revegetation of mine tailings is a challenging task in view of the fact that the metal mine tailings are usually very poor in nutrients, rich in toxic metal content, and have little capacity to retain water. Furthermore, wind erosion of mine tailings poses a serious environmental problem. All of these problems would be averted if tailings were revegetated with wetlands using metal-tolerant wetland plants. Wetlands have been used and constructed for the treatment of metal-con-

FIG. 8 Diagrammatic sketch of a cascade model of constructed wetland for removal of contaminants using aquatic macrophytes, i.e., Typha spp. and Scirpus lacustris. (Redrawn based on 79.)

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FIG. 9 Generalized plan of the cross section of large-scale constructed wetland.

taminated water in recent years, indicating that wetland plants that hyperaccumulate metals can tolerate elevated metal concentrations (Fig. 10). Glyceria fluitans (floating sweetgrass) is an amphibious plant found growing in the tailings pond of an abandoned lead-zinc mine in Glendalough, County Wicklow (38). Greenhouse experiments demonstrated that G. fluitans could grow in sand culture treated with high zinc sulfate solution. Further research confirmed that two populations of G. fluitans, one from a metal-contaminated and the other from a noncontaminated site, could be grown successfully on mine tailings with a high zinc content. G. fluitans and two other wetland plants, Phragmites australis and Typha latifolia, have since been grown on both alkaline and acidic zinc mine tailings in field conditions under fertilized and nonfertilized conditions. Research findings obtained so far indicate that G. fluitans can be easily established on zinc mine tailings. It appears also to have a very low nutrient requirement, thus keeping fertilizer costs to a minimum during rehabilitation of mine tailings. Wetlands have been constructed in Ireland for the passive treatment of tailings water originating from a lead-zinc mine. Water originating from mine tailings (slags) is often characterized by high metal and sulfate concentrations in comparison with background levels. Conventional methodology of tailing water treatment involves chemical treatment, which is a costly procedure requiring intensive chemical and labor inputs. Therefore, more recently both constructed and natural wetlands have been utilized for metal removal and wastewater quality control. Wetlands with their

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FIG. 10 Plants exhibiting four strategies when exposed to elevated concentration of metals. Indicators: Plants in which uptake and translocation reflect metal concentration in interstetial water and show toxic symptoms. Excluders: Plants that restrict the uptake of toxic metals into shoot over a wide range of background concentrations. Accumulators: Plants in which uptake and translocation reflect metal concentration in interstitial water without showing toxic symptoms. Hyperaccumulators: Plants in which metal concentration is more than 0.1% in dry matter. (Redrawn based on 9.)

diversified macrophytes are known to retain substances such as metals from water passing through them. Aquatic macrophytes encompass many common weeds enable cost-effective treatment and remediation technologies for wastewaters contaminated with inorganics and organics (78–82). 7

ACID MINE DRAINAGE

Acid mine drainage (AMD) significantly impairs the quality of aquatic ecosystems. The exposure and oxidation of iron sulfide from coal mining results in AMD. AMD is a significant source of water pollution for the Appalachian region of the United States. There, estimates by the Bureau of Mines indicate that about 20,000 km of US. streams and rivers are impaired by AMD. Therefore, in recent years several low-cost preventive and passive technologies have been developed that utilize biological and natural chemical processes to remediate the contamined

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mine waters without chemical treatment. In this regard, constructed wetlands and its assemblage, anoxic limestone drains (ALDs), and successive alkalinityproducing systems (SAPS) have potential, and promising results have been obtained in several instances (83,84). 8

METAL TOXICITY AND TOLERANCE

Induction of phytochelatins is one of the principal mechanisms of metal detoxification in aquatic macrophytes, e.g., Hydrilla verticillata, Certophyllum demersum, and Vallisneria americana. The induction of phytochelatins concomitantly reduced the cellular glutathione (43–51). The well-established mechanisms of metal toxicity and tolerance are depicted in Fig. 11 (85–90).

FIG. 11 Sources of metals and metal speciation in aquatic ecosystem with concomitant manifestations of toxicity and tolerance. Plant sources of heavy metals are mining (smelting, river dredging, mine spoils and tailings, metal industries, etc.), industry (plastics, textiles, microelectronics, wood preservatives, refineries, etc.), atmospheric deposition (urban refuse disposal, pyrometallurgical industries, automobile exhausts, fossil fuel combustion, etc.), agriculture (fertilizers, pesticides, etc.), and waste disposal (sewage sludge, leachate from land fill, etc.). Once metal ions are taken up by the cells, they are complexed in the cytosol with glutathione and the derived phytochelatins are transported into vacuole. Another notable cellular response is that some metals interact with genes having metal-regulating elements at the promotor region as found for metallothionein genes of animals (85–90).

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Methylcyclopentadienyl manganese tricarbonyl (MMT) has largely replaced tetraethyl lead as the antiknock compound in gasoline sold in North America. Manganese and lead uptake by plants growing near highways was much greater in aquatic plants than in terrestrial plants (Table 4) (91). Proximity to the roadway and volume of traffic greatly contributed to plant concentrations of metals. The Mn concentration increased over a growing season in Potamogeton pectinatus. By October the Mn levels exceeded levels known to be dangerous to a number of animals, including man (92). In June, the form in the leaves was Mn2⫹. By October it had shifted to predominately the Mn3⫹ form (93). In iron and manganese the Mn2⫹ form is mobile whereas the Mn3⫹ form is immobile. The concentration of Mn in aquatic plants is directly related to the density and proximity of human activity (Table 6) (94). Time of exposure to the metal is also an important factor (46). Plant tissues can accumulate very high levels of Mn and other metals before showing visible symptoms (95). Long before visible symptoms of toxicity appear, metabolic response to metals can be measured. Lemna minor L. treated with methyl mercury for just 12 h showed altered metabolism (Fig. 12) (94). Both respiration and photosynthesis were inhibited by concentrations of methyl mercury greater than 0.1 µg ml⫺1. At lower concentrations (0.0001, 0.001, and 0.01 µg ml⫺1) of methyl mercury respiration was stimulated to levels higher than control plants, probably due to mobilization of defenses against free radicals. Frond exposure to KCN and SHAM (salicylhydroxamic acid) under methyl mercury treatments revealed variation in the cytochrome oxidase pathway (Fig. 12). Low levels of methyl mercury stimulated the cytochrome pathway. Little change was noted in SHAM-sensitive respiration. 9

IRON PLAQUE AND ITS SIGNIFICANCE IN AQUATIC MACROPHYTES

Several macrophytes have shown the deposits of ion oxides or hydroxides on their root surface; these are called iron ‘‘iron plaque.’’ It is believed that wetland plants exhibiting iron plaque are expected to tolerate the metal-contaminated environment. Iron plaque acts as a barrier for the uptake of toxic metals that get adsorbed or immobilized by the plaque. The role of plaque in different species and geochemical settings of the sediments warrants further attention (96). In macrophytes of Bull Island, Dublin, Ireland, the anaerobic conditions caused iron plaque formation on roots of plants due to oxidation of iron in the rhizosphere. Using six different species of grasses and flowering plants, a comparison has been made on the degree of iron plaque formation. The techniques employed include light microscopy using JB-4 resin and staining with potassium ferrocyanide. It was observed that metals accumulated in the rhizosphere of wetlands and in the burrow walls of sediment infauna. Preliminary studies have suggested that salt marsh plants and sediment infauna could significantly affect the

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FIG. 12 Respiratory carbon dioxide from Lemna minor L. In the dark at different concentrations of methylmercury (MeHg) after 30-min exposures to the metabolic inhibitors potassium cyanide (KCN) and salicyl hydroxamic acid (SHAM), compared with control (94).

retention capacity of wetlands for metals through oxidation of the rhizosphere/ burrow wall. It was hypothesized that the thin oxidized soil zones around salt marsh-plant roots and invertebrate burrows provide a source of iron oxyhydroxides an otherwise anoxic, chemically reduced soil, which would result in the accumulation of various heavy metals and metalloids in the direction of the root/ burrow. At Bull Island salt marsh (Dublin, Ireland), Fe and As concentrations were significantly higher in the rhizosphere of the two plant species (Spartina townsendii and Atriplex portulacoides) and in the burrow wall of the lungworm Arenicola marina, compared with the adjacent bulk soil. Zn showed the same trend. A similar study, conducted at Rogerstown Estuary (Dublin, Ireland), found that both Zn and As accumulated in the rhizosphere of the same plants, but Fe did not. There are suggestions that the high organic matter content of this marsh may somehow interfere with oxidation processes involving Fe and/or with the binding of Zn and As to iron oxyhydroxides. More recent studies carried out at two marshes in South Carolina, USA, have shown no significant accumulation of Fe, Zn, or As in the rhizosphere of Spartina alterniflora (short and tall form)

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or in the burrow walls of Uca pugnax and Arenicola christata. It seems that the hypothesized accumulation processes that were observed at Bull Island cannot be assumed to occur universally. The study highlights the heterogeneous chemical nature of salt marsh soils, even to within a few millimeters—a factor that must be taken into consideration when sampling in such systems and when managing wetland ecosystems (97–99). 10 FOOD Some aquatic plants are used as food and feed. Water spinach (Ipomea aquatica) is commonly used as a vegetable and pig food in Thailand. It grows fast, easily, and is cultivated wherever there is water, as well as in industrial areas and big cities. It takes up and accumulates heavy metals to such an extent that its possible threat to human health has been discussed (17,70). Resent studies have shown

FIG. 13 Productivity (t ha⫺1 yr⫺1) of selected aquatic macrophytes under specific conditions that enable them to act as potential agents of phytoremediation.

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TABLE 7 Examples of Phytoremediation of Aquatic Ecosystems Using Macrophytes Location

Application

Metal(s)

Medium

Chevron Corporation Western USA

Phytoremediation Arsenic Photovolatilization Selenium

Wetlands Water

California, USA Stockholm, Sweden

Photovolatilization Selenium Phytoremediation Cadmium, copper lead, and zinc

Drainage, water Storm water ditch, sewage wetland

Ukraine uranium tailings Rhizofiltration

137

Cesium Strontium 226 Radium

Water

90

Elliot Lake, Canada

Phytoremediation

Watershed

Plant(s) used Typha, Nymphea (water lily) Variety of wetland plants and rhizosphere bacteria Salicornia Phlaris arundinaceae Typha latifolia Scirpus sylvaticus Hydroponic culture of crop plants Typha latifolia

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that plants collected in Thailand, close to Bangkok, contained up to 530, 350, and 123 (µg g⫺1 DW) of Pb, Hg, and Cd, respectively (Go¨thberg, Bengtsson, Greger, and Karlsson, unpublished). The biggest problem seems to be Hg. According to FAO, the intake of Hg shall not overcome 43 µg/week, which means that not more than 250 g of plant per day can be eaten. This plant has also been tested for removal of metals from wastewaters (100).

11 CONCLUSIONS Aquatic macrophytes listed above have several negative characteristics, i.e., hardiness, ability to survive under adverse environmental conditions, and high productivity (Fig. 13). This enables them to act as potential agents of phytoremediation (Tables 7 and 8). Furthermore, an anaerobic environment in the water and sediments renders elements in less oxidized forms. This increases the uptake by aquatic plants and may make it possible to effectively use aquatic plants in phytoremediation of wetlands. The partial pressure of oxygen in water is only a small fraction of the 21% oxygen in air; the aquatic environment itself is a strong defense against free radical formation. Unfortunately, low levels of oxygen also greatly reduce the availability of energy liberated by catabolism. This can hinder growth, particularly in stagnant ponds or other situations lacking natural aeration. The aquatic environment, especially bottom waters and sediments, often have a low oxygen content. Metals are more easily taken up by plants in reduced form. Once in the plant they may be oxidized and become immobile. Thus, aquatic macrophytes generally have a much higher metal content than terrestrial plants and can be usefully employed in phytoremediation. Aquatic macrophytes both living and nonliving serve not only as accumulators and indicators of pollution but also mitigate metal pollution to a considerable extent (101–112).

TABLE 8 Wetlands and Their Service for the Clean-up of Metal-Contanimated Aquatic Ecosystems Wheal Jane tin mine in Cornwall, UK Tennesse Valley Corporation, USA West Glamorgon, Wales, UK Ranger uranium mine in the northern territory, Australia Weir International Mining Consultants, West Virginia, USA Note: The above list is not exhaustive.

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ACKNOWLEDGMENTS One of the authors, MNVP, is grateful to the Department of Science and Technology, New Delhi (No. SP/SO/A21/97) dt 28-3-97, the Ministry of Environment and Forests, New Delhi, Govt. of India (No.19/33/95-RE) dt 28-3-97, Council of Scientific and Industrial Research, New Delhi (No. 38(901) 95/EMR-II dt 3010-1995.); the Third World Academy of Sciences, Italy (RGA No: 98-167 RG/ BIO/AS dt 21.12.98) and the Swedish International Development Agency, Sweden for providing assistance to deal with research in the area of ‘‘trace metals.’’ REFERENCES 1. MD Brasier. An outline history of seagrass communities. Palaeontology 18:681– 702, 1975. 2. L St-Cyr, PGC Campbell. Trace metals in submerged plants of the St. Lawrence river. Can J Bot 72:429–439, 1994. 3. L St-Cyr, PGC Campbell, K Guertin. Evaluation of the role of submerged plant beds in the metal budget of fluvial lake. Hydrobiologia 291:141–156, 1994. 4. JG Dean, FL Bosqui, VH Lanouette VH. Removing heavy metals from waste water. Environ Sci Technol 6:518–522, 1972. 5. U Fo¨rstner U, GTW Wittmann GTW, eds. Metal Pollution in the Aquatic Environment. Berlin: Springer-Verlag, 1979. 6. U Fo¨rstner. Metal transfer between solid and aqueous phases. In: Fo¨rstner U, Wittman GTW, eds. Metal Pollution in the Aquatic Environment. Berlin: SpringerVerlag, 1979, pp. 197–270. 7. M Greger, L Kautsky. Use of macrophytes for mapping bioavailable heavy metals in shallow coastal areas, Stockholm, Sweden. Appl Geochem (suppl.) 2:37–43, 1993. 8. BT Brown, BM Rattigan. Toxicity of soluble copper and other metal ions to Elodea canadensis. Environ Pollut 20:303–314, 1979. 9. Greger M. Metal availability and bioconcentration in Plants. In: MNV Prasad, J Hagemeyer, eds. Heavy Metal Stress in plants—From molecules to ecosystems. Berlin: Springer-Verlag, 1999, pp. 1–27. 10. MR Rattray. The relationship between P, Fe and Mn uptakes by submerged rooted angiosperms. Hydrobiologia 308:117–120, 1995. 11. RPH Welsh, P Denny. The uptake of lead and copper by submersed aquatic macrophytes in two english lakes. J Ecol 68:443–455, 1980. 12. G Chawla, J Singh, PN Viswanathan. Effect of pH and temperature on the uptake of cadmium by Lemna minor L. Bull Environ Contam Toxicol 47:84–90, 1991. 13. M Greger, L Kautsky. Uptake of heavy metals by macrophytes—a comparison between field samples and controlled experiments. In: E Bjornestad, K Jensen, eds. Proc. 12th Baltic Marine Biol. Symp. Fredenborg: Olsen & Olsen, 1992, pp. 67–69. 14. S Sinha, UN Rai, RD Tripathi, P Chandra. Chromium and manganese uptake by Hydrilla verticillata (L.f.) Royle: amelioration of chromium toxicity by manganese. J Environ Sci Health A 28:1545–1552, 1993.

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35.

36. 37.

38.

39.

40.

41.

42. 43.

44.

45.

46.

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10 Aluminum Toxicity in Acid Soils Plant Responses to Aluminum

Hideaki Matsumoto, Yoko Yamamoto, and S. Rama Devi Okayama University, Kurashiki, Japan

1 1.1

GENERAL ASPECTS OF ALUMINUM STRESS IN ACID SOILS Introduction to Acid Soils

Acidity is a major degradative factor of soils and covers an extensive area of both tropical and temperate regions (Fig. 1). Acid soils occupy nearly 30% (3.95 bha) of the arable land area (1) in both tropical and temperate belts. Depending on the degree of weathering and soil acidity, the world acid soils are classified into eight groups (entisols, inceptisols, andisols, spodosols, alfisols, ultisols, oxisols, and histosols). Among them, oxisols, ultisols, and alfisols represent the majority of the acid soils in the tropical region (1). The cold and temperate acid soils are mainly dominated by spodosols, alfisols, inceptisols, and histosols, and the tropical belt is dominated by ultisols and oxisols. Oxisols are characterized by having mainly the oxides of Fe and Al; they consist of clay and highly insoluble minerals, such as quartz. This soil condition 289

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FIG. 1 Distribution of acid soils in the world. (Courtesy of Kluwer Academic Publishers, London.)

represents the advanced stage of weathering and has a very low water-holding capacity. Ultisols are the most widespread type of soil in the warm humid regions. Kaolinite is the major predominant clay mineral with minor amounts of silicates. Alfisols are also the major predominant soils in the humid regions, with kaolinite as the predominant clay mineral with admixtures of small amounts of silicates. The factors involved in producing acid soils are the prolonged leaching by rainwater, soil-forming processes, and climatic conditions (2). In addition to these natural processes, even the agricultural farming processes and management practices, such as high use of nitrogen fertilizers, removal of cations by harvested crops, leaching and runoff of cations, resulted in lowering of soil pH (3). In many industrialized areas, the atmospheric deposition of sulfur and nitrogen compounds is a major source of proton influx to soils (2). More than the low pH of the soils, the major problem associated with acid soils is the toxicity of aluminum (Al) and manganese and the deficiency of phosphorus, calcium, magnesium, and potassium (4). In addition to these nutritional factors, the acid soils are also characterized by low water-holding capacity due to compaction of soils. Apart from mineral toxicities, soil acidification is also known to change the species spectrum of the forest soils by changing the microbial activity of the soils (5). Acidification is also known to reduce the degradation of soil organic matter and alters cation and nutrient flow in the ecosystem. In general, most of the acid soils have low

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exchangeable bases (e.g., Ca, Mg, and K) mainly because of the low cation exchange capacity (CEC) of the soils (6). The oxides of Al and Fe in wet acidic soils fix a large fraction of the phosphate, making it unavailable to plants leading to lower crop yields. Therefore, Al toxicity is the major agronomic problem in acid soils. Since Al is the major component of acid soils, this chapter mainly focuses on the understanding of Al toxicity and tolerance mechanisms in plants, which would be highly useful for selecting and breeding plants for Al resistance and to reclaim acid soils. 1.2

Aluminum Toxicity in Acid Soils

Al is the most abundant element in Earth’s crust, and globally soils with Al problems cover up to 67% of the total acidic soils. In acidic soils, the phytotoxic species of Al is solubilized to the levels that inhibit root growth and in turn increase crop losses. Because of its agronomic importance, much research has been focused on basic and applied aspects of Al toxicity and tolerance. Recently, considerable effort has been made to understand the Al toxicity and tolerance at whole-plant, cellular, subcellular, and molecular levels (7,8). 1.2.1

Chemistry of Aluminum

The chemistry of Al in water and soil is very complex as it exists in various forms depending on the pH of the solution and its complex interactions with soil and plant system (9). In the soil, Al exists in various forms, e.g., hydroxyoxides, aluminosilicates, and aluminophosphates (10). Under acidic conditions, the toxic Al species are released into solution. The released Al in soil solution can be phytotoxic and poses a major environmental risk (11). There is a lot of ambiguity about the toxic forms of Al that inhibit the plant growth. In solution, Al forms numerous complexes with organic and inorganic ligands (12). It is assumed that monomeric Al3⫹ is more toxic than hydroxy mononuclear forms Al(OH)2⫹ and Al (OH)2⫹ (13). Moreover, recently it has been demonstrated that under a certain low pH, polynuclear Al, such as Al13 can also exist in solution, which is much more toxic than the monomeric Al3⫹ form (14). However, the possibility of the existence of such polynuclear forms in soil solution is thought to be very low (9). 1.2.2

Sites of Aluminum Toxicity in Plant Cell

In addition to its complex chemistry, there is controversy with regard to Al toxicity sites in plants (11,15). However, several reports reveal that Al is largely associated with cell walls and plasma membrane because of negative charges of the cell wall pectin and phospholipids of the plasma membrane (7,16). Al toxicity mainly depends on the Al concentration and on the duration of treatment and plant

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species. However, whether the toxic effect of Al is exhibited in the apoplasm or symplasm is still not clear. In spite of many studies on the sites for Al toxicity, no concurrent evidence for any one particular site specific to Al toxicity has been known. 1.3

Aluminum Toxicity in Plants

The toxic effects of aluminum include stunted root growth, poor root hair development, swollen root apices, and stubby and brittle roots resulting in the decrease in translocation of water, nutrient elements, and phosphate from roots to shoots (7,8). However, there exists a wide variation among the species or the cultivars of the same species with regard to Al tolerance. The primary toxic effect of Al is the inhibition of root growth. In most situations, Al appears to be toxic to actively dividing and elongating cells and not to mature cells (17). This has been demonstrated with many plant species. In Al-sensitive maize cultivar, it has been shown that the transition zone of the root apex where cells undergo rapid division and elongation is more sensitive to Al toxicity than other parts of the roots (18). Blancaflor et al. (19) also reported that Al toxicity is associated with elongation zone of maize roots. In addition to roots, suspension-cultured tobacco cells are more sensitive to Al toxicity at the actively dividing log phase than the stationary phase, indicating that actively dividing cells are the targets for Al toxicity (20). However, the cytotoxicity varied widely depending on the concentration and duration of Al treatment. In Al-sensitive plants, Al at concentrations as low as 10 µM inhibits root growth within 10min after the Al application at pH 4.5 (21). Al exists mostly as Al3⫹ at this pH, but other forms of Al such as Al(OH)2⫹ and Al (OH)2⫹ also exist. However, the short-term results of Al toxicity mostly correlated with monomeric forms of Al. As the effect of Al varies with the concentration and duration of the treatment, Al is assumed to have multiple sites of action. Moreover, whether these effects are primary or secondary responses is difficult to analyze. Since root growth inhibition is the immediate response of the plants to Al, Al toxicity seems to be external or apoplastic in nature (mainly the cell wall and outer portion of plasma membrane). One of the proposed mechanisms for the toxicity of Al is Al-mediated rapid and irreversible displacement of Ca from cell walls, especially Ca bound to pectin. Because of the greater affinity of Al to pectin compared with Ca, it is likely that the binding of Al makes the cell wall more rigid and prevents its loosening, which is essential for cell expansion (21,22). It is also known that Al alters the cellular Ca homeostasis and the blockage of Ca channels of the plasma membranes (23). In tobacco BY-2 cells, the growth inhibition by Al is linked to Al-induced decrease in cytosolic Ca levels (23). In addition to cell walls, Al is known to displace Ca from the bridges between phospholipid head groups, causing increased phospholipid packing due to its small size and trivalent charge.

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In addition to displacement, Al is also known to lower the uptake of calcium by cells (24–27). Since Ca is a key element involved in the signaling process, the change in Ca levels would affect various metabolic processes, which are linked to Ca-dependent signal transduction (25,28). Another important phenomenon associated with Al toxicity is the accumulation of callose. This has been considered as one of the sensitive markers of Al toxicity. Several studies have shown that callose (β-1,3-glucane) accumulation starts immediately after the exposure to Al, and a strong correlation between Al toxicity and callose accumulation has been observed (29). However, in Al-tolerant Arabidopsis mutants, no obligatory relationship has been observed between Al uptake and callose accumulation (30). Although several studies support the idea that a major portion of Al is associated with cell walls or apoplast, recent studies with gaint cells of Chara indicated that Al is able to pass through the plasma membrane and enter into symplast (31). In addition to cell wall pectin, another potential site of Al binding is the phospholipids of the plasma membrane (32). In spite of a few attempts to understand the Al effects on membrane functions, no conclusive evidence has been obtained. Al-mediated changes in the lipid composition have been further demonstrated by Zhang et al. (33) who observed genotypic-dependent changes in lipid composition using Al-sensitive and tolerant plants. However, it is still uncertain whether there is any relationship between Al toxicity and the changes in lipid composition. In addition to the changes in lipid composition, Al is also known to induce some membrane proteins (34). However, the implication of their role in Al tolerance is not clear. 1.4

Aluminum Tolerance in Plants

There has been considerable variation among the plant species in Al tolerance. This is due to the variations in the accumulation of Al and the tolerance to the toxic effects of Al. Al tolerance in plants is achieved by either avoidance or internal tolerance (7,8). The plants acquire Al tolerance simply by avoiding the metal accumulation. This has been achieved by exuding organic acids such as malate, citrate, and oxalate from the roots, which chelate the toxic Al rendering it nontoxic to plant (35–37). On the other hand, plants developed some internal resistance mechanism(s) against Al. The factors that favor Al exclusion include (a) release of organic ligands such as malate, citrate, oxalate; (b) increase of rhizospheric pH by roots; (c) increased binding to cell walls; (d) decreased permeability of plasma membranes to Al influx; and, (e) binding of Al in the mucilage. Among them, there is increasing evidence for the chelation of Al by organic acids. This is evident from the studies showing that the external application of organic acids protects the plants from Al toxicity. The exclusion of organic acids by Al has been demonstrated in several plants. However, the type of organic acid

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secreted varies with the plants. The high resistance of buckwheat plants to Al has been ascribed to internal detoxification as well as rapid exclusion of oxalic acid from the roots (38). In wheat, malate exudation is important for the chelation of Al (39). Using near-isogenic lines, Delhaize et al. (35) convincingly demonstrated the role of organic acid excretion in Al tolerance. In maize, Pellet et al. (40) demonstrated the release of citrate from the Al-tolerant cultivar. These results suggest that organic acid exudation play a major role in Al tolerance. However, studies with wheat cultivars showed that malate exudation did not always correlate with the root growth in the presence of Al, suggesting that there could be other resistance mechanisms operating in conferring Al tolerance. For instance, in Al-tolerant Arabidopsis mutants, Al is known to increase the rhizospheric pH by increased influx of H⫹ ions at the root tip region (41). Degenhardt et al. (42) provided the first evidence in support of the role of increased rhizospheric pH in Al tolerance using Arabidopsis mutants. In addition to rhizospheric pH change, some of the mutants of the same plant released organic acids in response to the exposure to Al, indicating that more than one tolerance mechanism exist in plants (41). In addition to these metal exclusion processes, it has been postulated that Al binding to mucilage could play an important role in Al resistance (43). Although more information is being added to the literature, there is no convincing evidence to support any of these hypotheses. However, with recent technical advances in dealing with Al, satisfactory insight into Al toxicity has been gained. The mechanism of Al toxicity and tolerance are discussed in detail in subsequent sections of this chapter. 2 2.1

PLANT RESPONSES TO ALUMINUM STRESS: THE WHOLE PLANT General Aspects of Plant Growth in Acid Soil

Aluminum (Al) is the most abundant metal in soil, comprising approximately 7%. A genotypical difference in Al toxicity or tolerance exists among plant species, and the growth of plants in acid soil is better in the tolerant genotypes. Therefore, the effect of acid soil on plant growth is strongly related to the toxicity of Al. Al has been known to inhibit the root growth of barley and rye in acid soil since 1918 (44). Soil acidity seems to be gradually increasing due to environmental problems, such as acid rain. Furthermore, the anticipated burst of the global population in the future has drawn attention to Al toxicity and tolerance in acid soil in the past few decades (8,45,46). 2.2

Inhibition of Root Elongation by Aluminum

Inhibition of root elongation is the first visible symptom under Al stress. Inhibition of root elongation in Al-sensitive maize occurred within 30 min (47). In

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most plant species, root elongation is markedly inhibited by Al at a micromolar level at low pH such as 4.5 in a simple solution containing Ca2⫹ alone. However, the rate of inhibition is decreased by coexistence of other ions. A change of the negativity of the root surface upon the association with cations lowers the accessibility of Al to the root. The inhibition of root elongation by AlCl3 in Al-tolerant (Atlas 66) and Al-sensitive (Scout 66) wheat is shown in Fig. 2. Root elongation in Scout 66 was apparently inhibited by a 3-h treatment with 5 µM Al but that of Atlas 66 was not inhibited up to almost 10 times higher concentration (48). The inhibition of root elongation was accompanied by a decrease of cell viability, which was detected by staining with propidium iodide, due to the damage of plasma membrane under Al stress. The decrease of cell viability coincided with the suppression of root elongation (49). 2.2.1

Localized Accumulation of Aluminum in Roots

The region of Al accumulation and inhibition of root elongation has been suggested to be restricted to the root apex including the root cap, meristem, and elongation zone (7,50). When Al-tolerant (Atlas 66) and Al-sensitive (Scout 66) wheat were treated with 50 µM Al and stained with hematoxylin, only the apices of root were stained in both cultivars. In Atlas 66, the root apex exposed to 20 µM Al for 48 h was strongly stained but that exposed to 5 µM Al was only faintly stained. On the contrary, the root apex of Scout 66 exposed to 5 µM was stained strongly (49) (Fig. 3). These results indicate that the amount of Al accumulated in the growing root tissue is related to the sensitivity of the plant to Al and that tolerant wheat has a specific mechanism to excrete Al from the root apex. 2.2.2

Morphological Change of Cells Induced by Aluminum Toxicity

Since the shortening of roots caused by Al is accompanied by an increase in the diameter of the root tip, the diameter of the second layer cells of the cortex in Atlas 66 was measured. In control roots, the cells in the portion 0.6–3.2 mm from the root cap junction are markedly long and small in diameter (Fig. 4). This indicates that this portion is the elongation zone. Al treatment inhibited the increase in the cell length but increased the diameter of root cells (50). The ratio of length to diameter of cells in the control root was three to four times larger than that in the Al-treated roots. In particular, the cells in the second and third layers of the cortex were highly swollen laterally. The swollen cells were characterized by the drastic accumulation of lignin on their cell wall. These results indicate that the root zone where the elongation was inhibited by Al is the zone where Al is accumulated. Recently, it was found that the distal part of the transition zone of the root apex, where the cells are undergoing a preparatory phase for rapid elongation, is the primary target of Al in corn (18).

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FIG. 2 Seedlings of Al-tolerant (Atlas 66) and Al-sensitive (Scout 66) wheat. Seedlings were grown in the presence or absence of Al (pH 4.5) in 100 µM CaCl2 solution for 5 days. Top (Scout 66): 0, 1, 2, 5, and 10 µM Al from left; bottom (Atlas 66): 0, 5, 10, 20, and 50 µM Al. (From Ref. 48.)

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FIG. 3 Hematoxylin-stained roots of Atlas 66 and Scout 66. Seedlings were grown in the presence or absence of Al for 48 h. (A) Atlas 66 treated with 0, 5, and 20 µM Al from the left. (B) Scout 66 treated with 0, 1, and 5 µM Al from the left. Bar indicates 1.0 mm. (From Ref. 49.)

2.2.3

Inhibition of Cell Division by Aluminum

Cell division in the meristematic zone of the root measured by counting the metaphase cells is inhibited by Al (51,52). Compared with the time required for the inhibition of root elongation, a longer time is needed for the inhibition of cell division under Al stress. This means that the primary action of Al is the inhibition of root elongation. The contribution of cell division is only 1–2% of the root elongation. However, inhibition of root elongation is substantially caused by the inhibition of the elongation of the cells in the elongation zone, which is derived from the adjacent meristematic zone; Al toxicity kills the plant root by inhibiting cell division. Thus, the binding of Al to DNA or chromatin that induces the condensation of their structure concomitant with decreased template activity may cause the inhibition of cell division (53–55). Additional evidence for the inhibition of cell division is the disruption of phragmoplast and spindle microtubules

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FIG. 4 Effects of Al on the length and diameter and the ratio of length to diameter of root cells in the second layer from surface in Atlas 66. Roots were treated with or without 20 µM Al for 24 h. Data are means of five or six samples (⫾ SE). (From Ref. 50.)

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in the logarithmic growth cells under Al stress (56). Also, the observation of the extrusion of nucleolor materials from the nucleus to cytoplasm supports the above speculation (57). 2.3 2.3.1

Physiology of Aluminum Toxicity Aluminum Toxicity and Cell Wall

It is important to elucidate the localization of the absorbed Al in the cell in order to understand the toxicity of Al. The major portion ranging from 30% to 90% of the absorbed Al is found in the apoplast (58). Externally applied Al binds to the binding site in the apoplast immediately. A significant correlation between the amounts of Al and phosphate was found in the cell wall. Thus, Al phosphate in the cell wall may act as a barrier for the transport of Al into the cytosol. Pectin has been proposed to be a candidate of the cell wall component to which Al binds (58,59). However, research on the binding of Al to pectin in vivo is scanty (60). Blamey et al. showed that the addition of Al to an artificial Ca pectate membrane reduced water permeability (59). This may be related to the change of topology of apoplast in vivo, which may regulate the apoplastic transport. 2.3.2

Callose Formation

Callose formation on plasma membrane is a specific phenomenon under Al stress. Synthesized callose is deposited in the apoplast. Supply of 50 µM Al for only 30 min induced the callose formation in the root apex in soybean seedlings (61). When roots were exposed to Al at a concentration greater than 10 µM, the callose concentration in the apical 10–30 mm of the root tip was inversely correlated with the root elongation in cow pea. Therefore, callose formation is related to a potent parameter for Al toxicity. Recently, callose was found to accumulate in plasmodesmata in wheat root, causing the inhibition of cell-to-cell trafficking of molecules through plasmodesmata (M. Sivaguru et al., personal communication, 2000). 2.3.3

Plasma Membrane

The plasma membrane is the first candidate for the target of Al because of its outermost location in the cell and high content of phosphorus as phospholipid. Al has a 560-fold higher affinity for the phosphatidylcholine surface than Ca2⫹ (62). Al3⫹ at 5 ⫻ 10⫺6 M can neutralize the negative surface charge of the plasma membrane and cause a shift of ζ potential of the membrane surface from ⫺30 mV to ⫹11 mV. Interestingly, the depolarization of the surface charge (ζ potential) occurred mostly at the root apex where Al was also localized the most in Altreated squash (Ahn et al., personal communication, 2001). In Al-tolerant wheat cultivar Dade, the membrane potential was rapidly and significantly depolarized by exposure to Al at high concentrations. The maximum depolarization was 150

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mV at 150 µM AlCl3. In contrast, the membrane potential of the Al-sensitive cultivar Romano was depolarized only slightly by Al (63). Binding of Al to plasma membrane caused the decrease of membrane fluidity (64). Lipid peroxidation in the root tips of soybean was enhanced only after a long-term treatment with Al (65). This lipid peroxidation was intensified by Fe2⫹. The enhanced lipid peroxidation by oxygen free radicals is a consequence of the action of Al on membrane structure. The physiological function of plasma membrane is also affected by Al. Plasma membrane H⫹-ATPase activity was inhibited concomitant with depolarization of the membrane surface potential in root apex in Al-treated squash (Ahn et al., personal communication, 2001). In cortex cells in the intact roots of Northern red val (Quercus rubra), Al significantly altered the activation energy required to transport water (⫹72%), urea (⫹9%), and monoethyl urea (⫺7%) across the cell membrane. Ion transport process was also altered by Al. At 100 µM Al, the influx of Ca2⫹, NH4⫹ and K⫹ was inhibited by 69%, 40%, and 13%, respectively, but the flux of NO3⫺ and PO43⫺ was enhanced by 44% and 17%, respectively (66). This is caused by the Al-induced positive charge of the membrane. Furthermore, Al blocked the inward-rectifying K⫹ channel. Contrary to plasma membrane, the effect of Al on tonoplast is scarcely investigated. In barley root, the ATP- and PPi-dependent H⫹ transport activity of the tonoplast vesicle is increased by Al stress. The defense mechanism of barley may be related to the maintenance of H⫹ homeostasis in the cytoplasm through the sequestration of Al into the vacuole by the activation of H⫹ transport activity (67). 2.3.4

Calcium

Much work has been done on Al toxicity in terms of Ca metabolism (68). Ca2⫹ influx was rapidly inhibited at the root apex in correlation with the inhibition of root growth in wheat (21). It is interesting that the inhibitory effect of Al on Ca2⫹ transport is stronger in Al-sensitive wheat (Scout 66) than in Al-tolerant wheat. The Al-induced inhibition of Ca2⫹ uptake was rapidly recovered by the removal of Al from the solution. Not only influx but also translocation of Ca2⫹ from root apex to the basal part was similarly repressed by Al (69). The antagonistic effect of the large amount of Ca2⫹ on Al toxicity is known (70) (Fig. 5). Al increased membrane permeability to nonelectrolytes such as urea but decreased the membrane permeability to lipids and water. The effects of Ca2⫹ on the permeability were opposite those of Al3⫹ (71). One of the plausible inhibitory mechanisms of Al on Ca2⫹ metabolism is the displacement of apoplasmic and membrane Ca2⫹ by Al. In the isolated cell wall equilibrated in 50 µM Ca2⫹ at pH 4.4, 100 µM Al displaced more than 80% of the Ca2⫹ with a half-time of 25 min (72). Recently, several researchers cast doubt on the disorder of Ca2⫹ metabolism by Al. Treatment of wheat with Al severely inhibited root growth but not Ca2⫹ uptake. Some ameliorating cations that severely inhibit root growth depress Ca2⫹ uptake in the presence of Al but not in its absence. The conclusion of Kinraide

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FIG. 5 Ameliorating effect of Ca2⫹ on the inhibition of pea root elongation caused by Al. Three-day-old pea seedlings were treated with 5 µM AlCl3 and CaCl2 at various concentrations for 24 h. Each point is the mean of results from five samples in each case. (䊐): control, (■): Al treatment. (From Ref. 70.)

et al. was that the Ca displacement hypothesis fails to explain Al toxicity, and that amelioration by cations occurs because of decreased membrane surface negativity and the consequent decrease in Al3⫹ activity on the membrane surface (73). Furthermore, an increase in cytoplasmic free Ca2⫹ or [Ca2⫹]cyt, was greater in Alsensitive wheat than Al-tolerant wheat, in spite of the general understanding that Ca2⫹ influx is inhibited by Al. The Al-related increase of [Ca2⫹]cyt was correlated with the inhibition of root growth and Al-induced increase in [Ca2⫹]cyt was recovered by removing Al. This disturbance of [Ca2⫹]cyt homeostasis can be one of the mechanisms of Al toxicity (74). However, there is still a discrepancy in the involvement of Ca2⫹ metabolism in Al toxicity, and more work is needed to clarify the feature of Al-Ca interaction. 2.4 2.4.1

Aluminum Tolerance Genes Responding to Aluminum

Unlike animals, higher plants are usually fixed in the soil by their roots and thus cannot escape from the stress. Therefore, plants have been forced to develop a

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specific way to avoid stress in the soil. Al-tolerant genes have been searched at different levels. At the chromosomal level, genes for Al tolerance in the mediumtolerant wheat variety ‘‘Chinese Spring’’ were found to be localized on chromosome arms 6AL, 7AS, 2DL, 3DL, 4DL, and 4BL, and on chromosome 7D (75). Major genes for the tolerance in rye seems to be located on 3R and 6RS. In Altolerant Atlas 66 wheat, D genome determined the tolerance to acid soil and consequently contributed to the increased adaptation of the hexaploid wheats during their evolution. However, not all genes for Al tolerance in Atlas 66 are located on the D-genome chromosome (76). From the molecular research, genes encoding glutathione S-transferase, peroxidase, blue copper-binding protein, phenylalanine ammonia lyase, 1,3-β-glucanase, cysteine proteinase, etc., were found to be induced by Al stress (77–79). However, various materials and experimental conditions were used in these studies, and further research is needed to evaluate whether these Al-induced genes are relevant to the mechanism of tolerance from an agricultural point of view. So far many mechanisms are proposed for the tolerance of plant against Al toxicity. Studies on the mechanism of exclusion of Al from the rhizosphere which represses Al toxicity may be interesting. 2.4.2

pH Regulation in the Rhizosphere

The solubility of Al depends on the solution pH. The most toxic Al3⫹ is solubilized at a pH lower than 4.5. Therefore, plants can reduce Al toxicity by increasing the pH in the rhizosphere, resulting in the formation of less toxic Al. The Altolerant Arabidopsis mutant alr-104 alkalinized the rhizosphere in the presence of Al but the wild type did not. Alr-104 induced a twofold increase in net H⫹ influx only at the root tip, causing a rise in the pH at the root surface by 0.15 unit. Thus, the Al resistance in alr-104 is caused by the Al-induced pH increase in the rhizosphere. In the absence of Al, no difference in root H⫹ fluxes was observed between wild type and alr-104 (42). 2.4.3

Mucilage

Al forms a chelate complex with ligands and becomes less toxic. Mucilage and organic acids are considered to be natural Al ligands that are excreted from the roots. Root apices of most plant species are covered with a mucilaginous substance that is exuded from root cap cells. Mucilage consists mainly of polysaccharides containing uronic acid and has a high Al binding capacity. Fifty percent of the total Al of root apices was associated with mucilage in cowpea (80). Al bound to mucilage accounted for approximately 25–35% of the Al remaining after desorption in citric acid solution in the wheat root. It has been postulated that the binding of Al to mucilage is one of the Al-tolerant mechanisms. This was supported by the demonstration that removal of mucilage prior to treatment with Al facilitated accumulation of Al in root apices and enhanced Al rhizotoxicity in cowpea (80). On the contrary, mucilage of Zea mays was not effective in pre-

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venting Al-induced root inhibition in hydroponic culture. The total binding capacity of the mucilage was too small, although mucilage strongly binds Al and the bound Al was not phytotoxic (81). 2.4.4

Organic Acids

The most powerful candidate for the natural Al-chelating substance is an organic acid such as citrate, malate, and oxalic acid. The first related evidence is the excretion of citrate from snapbean root under Al stress. The root of the Al-resistant cultivar excreted 70 times more citrate in the presence of Al than in the absence of Al and 10 times more citrate than the Al-sensitive cultivars (82). Altolerant wheat cultivars excreted 5- to 10-fold more malic acid instantly after exposure to Al stress. The root apex is the primary region of malic acid excretion, and the excretion is specifically induced by Al not by other trivalent cations. Organic acids are supposed to be excreted through the anion channel of the plasma membrane. Buckwheat, which is tolerant to acid soil, excretes oxalic acid promptly after exposure to Al (83). Taro also excretes oxalic acid. The mixture of citric acid and Al at a molar ratio higher than 1 : 1 did not inhibit the root elongation in wheat (Fig. 6). The excretion of citrate is a rather slow process in comparison with that of malate and oxalic acid and needs the biosynthesis by increased activity of citrate synthase. In this respect, it was recently reported that the transgenic plants of tobacco and papaya to which the citrate synthase (CS) gene from Pseudomonas aeruginosa was introduced together with the 35S promotor of calliflower mosaic virus was tolerant to Al. They accumulated a large amount of citrate and excreted it from the root (84). However, the regulation of the excretion of organic acids by the Al signal is important because continuous excretion of organic acids due to the overexpression of the genes may consume an extraordinarily large amount of energy and carbon. This regulative mechanism should be solved from a practical point of view. It is interesting to know why excretion of organic acid under Al stress is different among plant species. The effect of organic acid on Al detoxification depends on the stability constant of the complex of organic acid and Al. 2.5

Beneficial Effect of Aluminum

Some plants are known to grow better in the presence of Al. The beautiful blue color of hydrangea (Hydrangea macrophylla) is caused by the specific pigment, delphinidin diglycoside, which chelates with Al. Hydrangea can grow under high Al stress and contains a large amount of Al in the cell sap of leaves. Al was found to be chelated with citrate at a molar ratio of 1: 1 (Al/citrate) forming less toxic Al (85). Tea contains as high as 30,000 ppm Al in epidermal cells of old leaves but only 600 ppm in young leaves (86). Al is localized in epidermal cells of old leaves. The shape of epidermal cells varies considerably between old and

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FIG. 6 Effect of molar ratio of Al to citric acid on the root elongation of wheat, Scout 66. The Al concentration was 50 µM with 0.1 mM CaCl2 (pH 4.5). Seedlings were cultured in the solution containing Al and citric acid at various ratios for 9 h. Error bars represent ⫾SD (n ⫽ 3). (From Ref. 91.)

young leaves. The epidermal cells expand horizontally and the transversal length is greater than the longitudinal length in old leaves. The well-thickened cell wall is a feature of epidermal cells of old leaves. Al may be accumulated in the cell wall during thickening (Fig. 7). Also, the strategy of tea plants against Al toxicity is that most of Al is bound to catechine, although some Al binds to phenolics and organic acids (87). Al promotes the growth of tea, which is characterized by the stimulation of new root formation (Fig. 8). The mechanism of the promotive effect of Al is unknown, but one explanation is the regulation of phosphorus utilization by the formation of an Al-phosphate complex because the tea plant is sensitive to an excess amount of phosphorus, which is harmful for the growth of tea (88). Growth of Melastoma malabathricum, Melaleuca cajuputi, Acacia mangium, Hydrangea macrophylla, Vaccinium macrocarpon, Polygonum sachalinense, and Oryza sat-

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FIG. 7 Upper epidermal and palisade cells of old (left) and young (right) tea leaves, ⫻2800. Transverse sections were embedded in epoxy resin and stained with iodine. (From Ref. 86.)

iva are stimulated more or less by Al. Plants whose growth is stimulated by Al are classified according to the criteria of Al accumulation: (a) Al excluders such as M. cajuputi, A. mangium, O. sativa; (b) Al root accumulators such as V. macrocarpon and P. sachalinense; and (c) Al accumulators such as M. malabathricum and H. macrophylla (89). The mechanism of tolerance in H. macrophylla has been investigated intensively and the internal chelation of Al with citrate (mol ratio ⫽ 1 :1) was detected. However, the mechanism of growth stimulation by Al is not known. M. melabathricum, M. macrocarpon, and A. mangium are well adapted to low-pH soils, such as peat soils and acid sulfate soils in the lowland tropics of Thailand and Malaysia. V. macrocarpon and P. sachalinense are also well adapted to peat soil and low-pH soils in the temperate region. Growth of tropical plants is enhanced by Al and NH4 application. In all plant species, the pH of the culture solution is decreased and the concentration of soluble Al and P is increased by the application of NH4 , which promotes growth (90). Roots of M. cajuputi excrete a large amount of citrate, and the excretion is slightly increased by Al treatment. The mechanism by which Al benefits the growth of plants is

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FIG. 8 Tea seedlings grown in the cultural solution for 100 days with (right) or without (left) 10⫺3 M AlCl3. (From Ref. 86.)

not clear and may vary depending on the plant genotype and growth condition, especially mineral composition in the medium. 3 3.1

PLANT RESPONSES TO ALUMINUM STRESS: PLANT CELLS IN SUSPENSION CULTURE Cultured Plant Cells as a System to Analyze Cellular Metal Toxicity and Tolerance

Plant cells in suspension culture are useful to analyze responses to toxic metal ions (eg., Al, Na, Cd, Cu) at cellular and molecular levels. The physiological

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conditions of cells in suspension culture can be more easily adjusted than those of whole plants in water culture. Thus, it is possible to prepare a physiologically homogeneous cell population in large quantities. Furthermore, we can easily isolate cell lines exhibiting a metal-tolerant phenotype. However, it is not always possible to examine whether the tolerant phenotype selected is heritable or not, since cells in suspension culture gradually lose the capacity to regenerate intact plants. In addition, in suspension culture the cells are always shaken to supply oxygen, which is also a kind of physiological stress. In spite of these disadvantages, cultured plant cell systems are useful for supporting studies of whole-plant systems. Several research groups have successfully used cultured plant cells for physiological and genetic studies of Al toxicity and tolerance. Here we describe the general precautions for handling a suspension cell system for analyses of metal toxicity and tolerance. Then we describe the mechanisms of Al cytotoxicity and tolerance elucidated by use of suspension-cultured cells. 3.2 3.2.1

General Precautions for Handling Suspension-Cultured Cells for Stress Studies Cell Lines

Plant cells in suspension culture are often derived from the callus generated from plant tissues (92). At the beginning of the suspension culture, there may be a mixture of cells showing heterogeneous characters, but the cells may maintain the capacity to regenerate intact plants. Thus, the cells during the primary cultures should be used for the isolation of stress-tolerant cell lines, if the regeneration of intact plants from selected cells are planned (see Sec. 3.4.1). Instead, established cell lines, which were purified and have cultured for a long time without apparent changes of phenotype, seem to be composed of homogeneous cells. Thus, the reproducible responses to metal ions can be expected in the established cell lines, indicating that the established cell lines are useful for analyses of stress physiology (see Sec. 3.3) as well as for the isolation of stress-tolerant cell lines or genes (see Secs. 3.4.2 and 3.5). Several established cell lines from several species are available. We have used a tobacco cell line SL originally generated from pith callus of Nicotiana tobacum L. cv. Samsun (93,94). Another tobacco cell line, BY-2, is a kind of ‘‘standard cell line,’’ which has been widely used for physiological studies (95). 3.2.2

Cellular Growth Conditions

Plant cells in suspension culture are maintained by the transfer of portion of a fully grown cell suspension into new medium. In the case of tobacco cell line SL, cells do not grow just after the transfer (lag phase), begin to grow logarithmi-

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cally for a few days (log phase), and gradually stop growth (stationary phase) (20). Physiological conditions of cultured cells seem to change depending on the growth phase. Actually, Al sensitivity of tobacco cells was higher in a log phase than that in a stationary phase (20). Therefore, it is necessary to check the growth condition of cells before the start of experiments. The growth rate of cells can be roughly estimated by the population doubling time. 3.2.3

Treatment Medium

The metal ion of interest may interact with medium components and may affect plant cells differently depending on the treatment medium. Al ions have strong affinity with oxygen donor ligands such as phosphate (Pi) and carboxylate groups. Therefore, Al forms insoluble precipitates with Pi in medium and exhibits lower toxic effect. When the treatment time with Al is within 24 h, tobacco cells are not starved of Pi (96). Therefore, Pi can be withheld from the treatment medium (97). EDTA should be removed from the treatment medium for Al because it chelates Al (97). Other cationic ions may interfere with Al3⫹, since the ameliorative effect of cationic ions (eg., H⫹, Ca2⫹, Mg2⫹, K⫹, Na⫹) on the rhizotoxicity of Al3⫹ in wheat (Triticum aestivum L.) roots seems to be due to the reduction of the negativity of the cell surface electrical potential by the cations (73). Therefore, concentrations of cationic ions of treatment medium should be taken into consideration. Furthermore, it should be remembered that some metal ions interactively exhibit cytotoxicity. A combination of Al and Fe ions synergistically enhances lipid peroxidation in tobacco cells (see Sec. 3.3.2). Several metal ions change ionic species depending on medium pH. Below pH 5.0, Al ions are monomeric ions that can interact with cells and are toxic. Therefore, medium pH should be adjusted below 5. 3.2.4

Determinations of Viability

Cellular viability can be estimated quantitatively by (a) the uptake of Evans blue, (b) the formation of formazan from 3-(4,5-dimethylthiazol-2-yl)-2,5,-diphenyl tetrazolium bromide (MTT), or (c) the degree of posttreatment growth (98). Uptake of nonpermeable dyes, such as Evans blue, is caused by the loss of plasma membrane integrity, which is thought to be a marker of cell death (99,100). The formation of blue formazan from MTT is a marker of active cells, since MTT is a permeable dye and is reduced to form blue formazan by active mitochondria (101,102). The alteration of growth capability by metal treatment can be directly estimated by the relative growth of the treated cells to control cells during posttreatment culture (20,98). The general information for handling plant cells in suspension culture (92) and the experimental conditions for the study of Al stress using plant cells in suspension culture were also previously discussed (103,104).

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Mechanisms of Aluminum Cytotoxicity Responses of Cultured Tobacco Cells to Aluminum

3.3.1.1 Aluminum Accumulation The degree of the accumulation of Al in tobacco cells depends on medium components. In a simple medium containing 3 mM CaCl2 and 3% sucrose (pH 4.5), the accumulation of Al in cells initiated immediately after a start of Al exposure and increased until the most Al in medium accumulated in cells (105). This kinetic pattern was similarly observed in plant roots (106). In nutrient medium without Fe ions, the accumulation of Al was very limited. However, in nutrient medium with Fe ions, a burst of the accumulation of Al started after prolonged Al exposure (⬃8 h), which is due to the loss of integrity of the plasma membrane caused by the Al-enhanced Fe-mediated peroxidation of lipids (107) (see Sec. 3.3.2). It was also reported that the higher the CaCl2 concentration in medium, the lower the Al uptake in a tobacco cell line BY-2 in a simple medium containing 10 mM MES, 5 mM sucrose, 5 mM KCl, and CaCl2 (0.2, 2.0, 20 mM) (pH 4.5) (108). 3.3.1.2 Callose Production Callose synthesis is the most sensitive indicator of Al effect in plant roots (29,109). The Al-induced callose synthesis was identically detected in a simple calcium medium in both cultured soybean (Glycine max) cells (110) and tobacco cells (105). In soybean cells, the production of callose was detected as early as 15 min after the start of Al exposure. 3.3.1.3 Other Responses In Soybean cells, Al reduced net K⫹ efflux in a short-term exposure (110). The possibility that Al increases [Ca]cyt has not been fully examined because of the technical difficulty of detecting a transient change of [Ca]cyt (see Sec. 2.3.4). In a case of tobacco cell line BY-2, Al induced a decrease in [Ca]cyt , suggesting that Al may act to block Ca2⫹ channels at the plasma membrane (23). 3.3.2

Aluminum-Enhanced Peroxidation of Lipids

Al ions have a strong affinity for the surface of biomembranes (62), which causes the rigidification of the membrane (111). The Al-induced rigidification of the plasma membrane seems to affect the metabolism in the plasma membrane. Al stimulates the Fe(II)-mediated peroxidation of membrane lipids, and this nonenzymatic reaction was proposed to be due to the Al-induced rigidification of the membrane (112). The Al-enhanced peroxidation of lipids has been reported in various systems including phospholipid liposomes (112), soybean root tips (65,113), and cultured tobacco cells (114). In the cultured tobacco cells treated with Al in a simple calcium solution (see Sec. 3.3.1.1), the accumulation of Al in cells itself did not cause the peroxidation of lipids; but, the addition of Fe(II) to the Al-accumulated cells immediately resulted in lipid peroxidation, followed by loss of integrity of the plasma membrane several hours later (105). Thus, it

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seems that the accumulation of Al sensitizes the membrane to the Fe(II)-mediated peroxidation of lipids and that the Al-enhanced peroxidation of lipids is a direct cause of cell death (Fig. 9). The Al-enhanced peroxidation of lipids was also observed in pea (Pisum sativum L.) roots as an early response to Al without external supply of Fe ions (our unpublished results). The Al-enhanced peroxidation of lipids leads to the apoptosis-like cell death in cultured tobacco cells (115). In animal cells, cell death is a scheduled event (so-called programmed cell death or apoptosis) during a range of biological processes involving a relatively limited number of execution pathways. Internucleosomal fragmentation of DNA is the hallmark of apoptosis. In addition, an increase of intracellular Ca2⫹ and the activation of endogenous proteinases are known to participate in signaling pathways to promote apoptosis. Tobacco cells that had been treated with Al in the presence of Fe(II) in nutrient medium exhibited cell shrinkage and internucleosomal DNA fragmentation, which was detected as DNA ladders by agarose gel electrophoresis in the form of bands corresponding to increasing multiples of approximately 150-bp fragments (Fig. 10). Furthermore, the cell death process required extracellular Ca2⫹ and endogenous protein-

FIG. 9 Responses to Al of tobacco cells in Ca medium. The exposure of cells to Al in Ca medium leads to the accumulation of Al in cells and the deposition of callose, whereas the plasma membrane is apparently normal. The addition of Fe2⫹ to the Al-accumulated cells induces the Fe2⫹-mediated peroxidation of lipids, which results in leakage of K ions, enhancement of callose deposition, and loss of integrity of the plasma membrane (or cell death). (From Ref. 105.)

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FIG. 10 DNA fragmentation in tobacco cells treated with Al in the presence of Fe2⫹. Cells were treated with various concentrations of AlCl3 (0, 50, 100, 150 µM) in nutrient medium containing 100 µM FeSO4 (pH 4.0) for 24 h at a cell density of 10 mg fresh weight mL⫺1. After the treatment, protoplasts were prepared and nuclear DNA was isolated. DNA was analyzed by agarose gel electrophoresis followed by ethidium bromide staining. λ-Phage DNA digested with HindIII (designated as M) and 100-bp molecular ruler (indicated on the right side) were used for DNA markers. (From Ref. 115.)

ase. Thus, the cell process initiated by the Al-enhanced Fe(II)-mediated peroxidation of lipids is involved in the apoptosis-like cell death program. 3.4

Isolation and Characterization of Aluminum-Tolerant Cell Lines

Since responses to Al in cultured plant cells are similar to those in whole plant roots in water culture (see above), it is likely that some of the mechanisms of tolerance to Al are mediated at the cellular level. Based on this idea, several attempts have been carried out for the isolation of Al-tolerant cell lines under

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various screening conditions from various species (see below). The Al-tolerant phenotype expressed in some selected cell lines was elucidated to be due to genetic alterations (mutation) by the regeneration of the plants that maintained Al tolerance. The mechanism of Al tolerance was also investigated by use of Altolerant cell lines. 3.4.1

Isolation of Aluminum-Tolerant Cell Lines and Regeneration of Aluminum-Tolerant Plants

A large number of Al-tolerant variants were selected from nonmutagenized cell suspension cultures of Nicotiana plumbaginifolia Viv. (116). Furthermore, fertile plants were regenerated from the variants that retained stable Al resistance in callus culture, and all of them transmitted Al tolerance to their seedling progeny in segregation ratios expected for a single dominant mutation. Two types of Al-tolerant cell lines were selected from carrot (Daucus carota L.) cells in suspension culture (117). One type was tolerant to ionic Al, and the other was not tolerant to ionic Al but grew with insoluble Al-phosphate as a sole source of Pi. Both tolerant phenotypes were passed on from cells to the regenerated plants or plantlets (117). The reverse approach to getting Al-tolerant cells in suspension culture from Al-tolerant plant cultivars has not been successful. Cell suspensions were generated from the explants derived from an Al-sensitive (Romano) and an Al-tolerant cultivar (Dade) of Phaseolus vulgaris L. Although Al sensitivity was evaluated in these cultures using Al-induced callose synthesis, differences in callose production were not expressed between cell cultures (104). 3.4.2

Mechanisms of Aluminum Tolerance Expressed in Aluminum-Tolerant Cell Lines

In acid soils, the availability of Pi is reduced because of the precipitation of Alphosphate (45). The cell line selected from carrot cells (see Sec. 3.4.1) grew normally in Al-phosphate medium by secretion of citrate into medium. Citric acid chelates the Al in Al-phosphate, which results in the formation of nonphytotoxic Al-citrate and the solubilization of Pi. Furthermore, the cell line showed higher activity of mitochondrial citrate synthase (118), and the overexpression of mitochondrial citrate synthase gene of Arabidopsis thaliana in carrot cells improved the growth in Al-phosphate medium (119). Thus, the enhanced activity of mitochondrial citrate synthase seems to increase the productivity of citrate, which may increase cell growth in the presence of insoluble Pi source such as Al-phosphate. A tobacco cell line tolerant to the Al-enhanced peroxidation of lipids was isolated (120). Tobacco cells (a cell line SL) were mutagenized with ethyl methanesulfonate (EMS) to increase mutation frequency and then exposed to a lethal

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dose of Al in nutrient medium containing Fe ions. The survivors were grown in Al-free medium, and the Al-tolerant cell line was isolated. Compared to the parental cell line (SL), this cell line exhibited a 65-fold higher specific activity of phenylalanine ammonia lyase and contained 10-fold more phenylpropanoids. The phenylporpanoids (mainly caffeoylputresine) exhibited antioxidant activity against Al-enhanced peroxidation of lipids in the parental cell line (unpublished results). Thus, the phenylpropanoids that accumulated in the Al-tolerant cell line seem to protect cells from the Al-enhanced peroxidation of lipids. This cell line was also tolerant to the peroxidation of lipids caused by hydrogen peroxide, tertbutylated hydroperoxide, and A23187 (unpublished results). 3.5

Aluminum-Inducible Genes Isolated from Cultured Cells

Treatment with Al in the presence of Fe(II), which enhances the peroxidation of lipids in tobacco cells (see Sec. 3.3.2), induced genes for glutathione S-transferase, peroxidase, and GDP dissociation inhibitor (77,78,121). The possible involvement of these genes in Al tolerance was investigated by transferring these genes into Arabidopsis thaliana. All three genes conferred a degree of Al tolerance in A. thaliana, suggesting that these genes are involved in Al tolerance mechanisms (122). 4

SUMMARY

Suspension-cultured tobacco cells are useful for (a) investigation of the details of Al cytotoxicity at cellular and molecular levels, (b) isolation of a large variety of Al-tolerance cell lines, and (c) elucidation of possible mechanisms of Al tolerance and of possible Al tolerance genes. This system also seems to be useful for the study of toxicity and tolerance of other metals in plant cells. REFERENCES 1. VC Balliger, JL Ahlrichs. Nature and distribution of acid soils in the world. In RE Schaffert ed. Proceedings of the Workshop to Develop a Strategy for Collaborative Research and Dissemination of Technology in Sustainable Crop Production in Acid Savannas and Other Problem Soils of the World. Purdue University, 1998, pp. 1–11. 2. PA Sanchez, JR Benites. Low-input cropping for acid soils of the humid tropics: a transition technology between shifting and continuous cultivation. In: M Latham, P Ahn eds. Africaland: Land Development and Management of Acid Soils in Africa II. IBSRAM Proceedings No. 7, Bangkok, 1987, pp. 85–106. 3. B Ulrich. Production and consumption of hydrogen ions in the ecosphere In: TC Hutchinson and M Havans, eds. Effects of Acid Precipitation on Terrestrial Ecosystems. New York: Plenum Press, 1980, pp. 255–282.

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4. Von Uexkull, E Mutert. Global extent development and economic importance of acid soils. Plant Soil 171:1–5, 1995. 5. V Walters, RG Joergensen. Microbial turnover in beech forest soils at different stages of acidification. Soil Biol Biochem 23:897–902, 1991. 6. VC Balliger, NK Fageria. Nutrient use efficiency in acid soils at low pH: nutrient management and plant use efficiency In: AC Moniz, AMC Furlani, RE Schaffert, NK Fageria, CA Rosolem, H Cantarella, eds. Plant Soils Interaction at Low pH: Sustainable Agricultural and Forestry Production. Campinas, Brazil: Brazilian Soil Science Society, 1997, pp. 159–164. 7. E Delhaize, PR Ryan. Aluminum toxicity and tolerance in plants. Plant Physiol 107:315–321, 1995. 8. LV Kochian. Cellular mechanisms of aluminum toxicity and resistance in plants. Annu Rev Plant Physiol Plant Mol Biol 46:237–260, 1995. 9. S Funakawa, H Hirai, K Kyuma. Speciation of aluminum in soils solution from forest soils in northern Kyoto with special reference to their pedogenic process. Soil Sci Plant Nutr 39:281–290, 1993. 10. BS, Hemingway, G Sposito. Inorganic aluminum bearing solid phases In: G Sposito, ed. The Environmental Chemistry of Aluminum. Boca Raton: CRC Press, 1998, pp. 55–86. 11. LV Kochian, DL Jones. Aluminum toxicity and resistance in plants In: RA Yokel, MS Golub, eds. Research issues in aluminum toxicity, Washington, DC: Taylor and Francis, 1997, pp. 69–89. 12. WR Harris, G Berthon, JP Day, C Exley, TP Flaten, WF Forbes, T Kiss, C Orvig, PF Zatta. Speciation of aluminum in biological systems. In: RA Yokel, MS Golub, eds. Research Issues in Aluminum Toxicity. Washington, DC: Taylor & Francis, 1997, pp. 91–116. 13. TB Kinraide, DR Parker. Apparent phytotoxicity of mononuclear hydroxy aluminum to four dicotyledenous species. Physiol Plant 79:283–288, 1990. 14. DR Parker, TB Kinraide, LW Zelazny. On the phytotoxicity of polynuclear hydroxy-aluminum complexes. Soil Sci Soc Am J 53:789–779, 1989. 15. DL Jones, LV Kochian. Aluminum interaction with plasma membrane lipids and enzyme metal binding sites and its potential role in Al cytotoxicity. FEBS Lett 400: 51–57, 1997. 16. M Deleers, JP Servais, E Wulfert. Micromolar concentrations of Al3⫹ induce phase aggregation and dye release in phosphotidylserine-coatining lipid vesicles. Biochim Biophys Acta 813:195–200, 1985. 17. M Rincon, RA Gonzales. Aluminum partitioning in intact roots of aluminum-tolerant and aluminum-sensitive wheat (Triticum aestivum L.) cultivars. Plant Physiol 99:1021–1028, 1992. 18. M Sivaguru, WJ Horst. The distal part of the transition zone is the most aluminumsensitive apical root zone of maize. Plant Physiol 116:155–163, 1998. 19. EB Blancaflor, DL Jones, S Gilroy. Alterations in the cytoskeleton accompany aluminum-induced growth inhibition and morphological changes in primary roots of maize. Plant Physiol 118:159–172, 1998. 20. Y Yamamoto, S Rikiishi, Y-C Chang, K Ono, M Kasai, H Matsumoto. Quantitative estimation of Al toxicity in cultured tobacco cells: correlation between Al uptake and growth inhibition. Plant Cell Physiol 35:575–583, 1994.

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77. B Ezaki, Y Yamamoto, H Matsumoto. Cloning and sequencing of the cDNAs induced by aluminum treatment and Pi starvation in cultured tobacco cells. Physiol Plant 93:11–18, 1995. 78. B Ezaki, S Tsugita, H Matsumoto. Expression of a moderately anionic peroxidase is induced by aluminum treatment in tobacco cells: possible involvement of peroxidase isozymes in aluminum ion stress. Physiol Plant 96:21–28, 1996. 79. R Cruz-Ortega, JC Cushman, JD Ownby. cDNA clones encoding 1,3-β-glucanase and a fimbrin-like cytoskeletal protein are induced by Al toxicity in wheat roots. Plant Physiol 114:1453–1460, 1997. 80. WJ Horst, A Wagner, H Marschner. Mucilage protects root meristems from aluminum injury. Z Pflanzenphysiol 105:435–444, 1982. 81. XF Li, JF Ma, S Hiradate, H Matsumoto. Mucilage strongly binds aluminum but does not prevent roots from aluminum injury in Zea mays. Physiol Plant 108:111– 129, 2000. 82. SC Miyasaka, JG Buta, RK Howell, CD Foy. Mechanism of aluminum tolerance in snapbeans. Root exudation of citric acid. Plant Physiol 96:737–743, 1991. 83. JF Ma, SJ Zhang, H Matsumoto, S Hiradate. Detoxifying aluminum with buckwheat. Nature 390:569–570, 1997. 84. JM de la Fuente, V Ramirez-Rodriguez, JL Cabrera-Ponce, L Herrera-Estrella. Aluminum tolerance in transgenic plants by alteration of citrate synthesis. Science 276: 1566–1568, 1997. 85. JF Ma, S Nomoto, T Iwashita, H Matsumoto. Internal detoxification mechanism of Al in Hydrangea. Identification of Al form in the leaves. Plant Physiol 113: 1033–1039, 1997. 86. H Matsumoto, E Hirasawa, S Morimura, E Takahashi. Localization of aluminum in tea leaves. Plant Cell Physiol 17:627–631, 1976. 87. T Nagata, M Hayatsu, N Kosuge. Identification of aluminum forms in tea leaves by 27Al NMR. Phytochemistry 31:1215–1218, 1992. 88. S Konishi, S Miyamoto, T Taki. Stimulatory effects of aluminum on tea plants grown under low and high phosphorus supply. Soil Sci Plant Nutr 31:361–368, 1985. 89. M Osaki, T Watanabe, T Tadano. Beneficial effect of aluminum on growth of plants adapted to low pH soils. Soil Sci Plant Nutr 43:551–563, 1997. 90. T Watanabe, M Osaki, T Tadano. Effects of nitrogen source and aluminum on growth of tropical tree seedlings adapted to low pH soils. Soil Sci Plant Nutr 44: 655–666, 1998. 91. XF Li. Aluminum detoxification mechanism by organic acids and mucilage in higher plants. PhD dissertation, Okayama University, Kurashiki, Okayama, 2000. 92. DA Evans, WR Sharp, PV Ammirato, Y Yamada. Hand Book of Plant Cell Culture, Vol. 1. New York: Macmillan, 1983. 93. C Nakamura, M Nakata, M Shioji, H Ono. 2,4-D resistance in a tobacco cell culture variant: cross-resistance to auxins and uptake, efflux, and metabolism of 2,4-D. Plant Cell Physiol 26:271–280, 1985. 94. C Nakamura, HV Telgen, AM Mennes, H Ono, KR Libbenga. Correlation between auxin resistance and the lack of a membrane-bound auxin binding protein and a root-specific peroxidase in Nicotiana tabacum. Plant Physiol 88:845–849, 1988.

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11 Tree Crops Tracy Punshon Savannah River Ecology Laboratory, Aiken, South Carolina

1

INTRODUCTION

The preface to a recent text on biodiversity states that ‘‘the biota of the earth is being altered at an unprecedented rate. We are seeing massive changes in landscape use that are creating even more abundant successional patches, reduction in population sizes, and in the worst cases, losses of species’’ (1). A plethora of anthropogenic activities have forced changes in biodiversity, most notably changes in land use, unsustainable exploitation, and global pollution. Any organism unable to adapt to the new anthropogenic environments being created will either move or perish. Biodiversity is a measure of the number and abundance of species within an ecosystem and is therefore considered an indication of environmental quality (2). Natural forest ecosystems are believed to have the highest biodiversity, as they represent—in successional terms—a climax community; the last of many ecological seres. Environmental stresses that negatively affect growth and reproduction will reduce biodiversity, forcing out unadapted species by a process of either migration or extinction, while being balanced by an increase in biodiversity caused by the adaptation of other species to unique, localized contaminated niches (3). Much attention has been focused on the various environ321

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mental problems caused by widespread industrialization and urbanization, and the impact can be seen in every component of the biosphere. Global warming, caused by overproduction of CO 2 , is responsible for significant increases in temperature that impact species distribution. Combustion of coal for energy production releases large quantities of sulfur and nitrogen dioxides into the air, which combine with water to form acid rain, acidifying lakes and causing forest decline. Energy production from coal combustion also produces hundreds of millions of tons of ash waste per year [100 million tons in the United States alone (4)], the majority of which is stockpiled on land. A huge range of chemical waste products from various industrial and domestic sources is emitted in to the air, soil, and water. In an attempt to clarify how these various sources of contamination affect the biosphere they have generally been studied separately, although the effects observed on the wider environment are a result of highly complex interactions between contaminants and conditions, which is almost impossible to predict. However, the overall effect of all of these environmental stresses operating simultaneously is negative, possibly because the rate at which the environment is being altered is faster than the rate at which the majority of species can adapt. Continued selection pressure from chronic environmental contamination can prompt adaptational changes as well as toxicity. The survival of long-lived woody species depends on the ability to adapt in the short term to environmental change; and there is mounting evidence to suggest that some species may be able to take advantage of polluted environments and in some cases ameliorate them (5). Trees are now the focus of a great deal of research dealing with low-cost environmental clean-up, specifically for controlling hydrology, degrading organic contaminants in groundwater plumes, stabilizing eroded sediments, removing heavy metals from contaminated soil, and as biological filters for sewage sludge disposal. The research reviewed in this chapter focuses on adaptive changes of trees to heavy metal stress, with respect to their potentially beneficial uses for the remediation of heavy metals in the soil. The effects of metals on the physiological processes of woody plants, how these effects act together to produce changes on the whole ecosystem, and the mechanisms of adaptation of woody plants to anthropogenic selection pressures are also introduced. Several other highly successful remedial applications using trees have also been mentioned with the main aim of introducing the reader to the full potential of the use of trees for remediation, including degradation of organic chemicals, N and P removal, and wastewater purification, although they are left for others to review in detail. Their inclusion is intended to emphasize the potential of trees to remediate complex heterogeneous contamination rather than single-contaminant profiles, due to the predominance of mixed wastes. It is not the intention of this chapter to restrict the focus of the reader but rather to enlarge it.

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323

SIGNIFICANCE OF FOREST ECOSYSTEMS

Forest and related areas (e.g., woodland, shrubland, degraded forests, etc.) cover approximately 52 ⫻ 10 6 km 2 of the earth’s surface; an area three times that of cropland, and 75% more than grassland; estimated to be 20–45% of land cover (6,7). Forests are a vital part of the global ecosystem, producing 60% of the net primary productivity of all terrestrial environments and regulating water regimes by intercepting rainfall and regulating flow. They also play a vital role in enhancing soil quality, preventing soil erosion, adding organic matter from seasonal leaf fall, and enriching the diversity and abundance of soil microbes. It is not an exaggeration to say that the presence of woody vegetation has a profound ameliatory effect on the physical, chemical, and biological health of the soil. Tree crops form the basis of a wide variety of important industries, including timber, processed wood and paper, rubber, fruit, and coffee. Immense quantities of wood are produced annually; for example, in 1988 the total world timber production was conservatively estimated at 3.4 ⫻ 10 12 m 3; with 1.7 ⫻ 10 11 tonnes of paper produced that same year. In terms of international trade, more than 2.5 ⫻ 10 11 m 3 of wood and 5.1 ⫻ 10 10 m 3 of paper was traded in 1988 alone, amounting to $9 ⫻ 10 11, or 3% of world trade income (7).

3

SIGNIFICANCE OF HEAVY METAL CONTAMINATION

In the past, the term ‘‘heavy metal’’ referred to metals in the periodic table with an atomic number greater than 20, excluding the alkali and alkali earth metals (8), or metals that have specific gravities greater than 5 g cm ⫺3. The term is somewhat inaccurate, however, and alternative nomenclature has been suggested, with some workers preferring the term ‘‘potentially toxic elements’’ (9). The elements Cd, Cr, Cu, Hg, Ni, Pb, and Zn are generally considered the most important elements associated with soil pollution. Heavy metals occur naturally in rock formations and ore minerals (Table 1); therefore, they have a natural background concentration in soil, sediments, water, plants, and animals (10). Soils develop as a result of the gradual weathering of rock formations and typically inherit the characteristic metal assemblage of the parent rock. Igneous parent rock potentially contributes higher concentrations of Cr, Mn, Co, and Ni to the soils, whereas sedimentary parent materials (especially shales) contribute Cr, Co, Ni, Zn, and Pb (11). One of the most interesting examples of this phenomenon can be found in New Caledonia where serpentine rock formations, rich in Ni, Co, and Cr, give rise to soils with such elevated concentrations of metals that they are populated only with specifically adapted flora. These plants, the majority of which belong to the genera Alyssum and Thlaspi, have evolved a constitutional tolerance to the metals, and accumulate

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TABLE 1 Sources (Mineral and Anthropogenic) and Characteristic Information on the Heavy Metals Source Element Cadmium (Cd)

Chromium (Cr)

Copper (Cu)

Mercury (Hg)

Nickel (Ni)

Lead (Pb)

Zinc (Zn)

Mineral

Anthropogenic

Black shales: greenockite (CdS: ⬎77% Cd); otavite (CdCO 3 65.2% Cd) Carlsbergite (CrN ⬎ 78% Cr)

Phosphatic fertilizers, spoil heaps, sewage sludge disposal on land Metal finishing wastes, wood preservatives, sewage sludge Mining and smelting, metal processing (wire manufacturing), sewage Smelting processes, spoil heaps, sewage sludge disposal Iron, steel and pyrometallurgical industry wastes

Chalcocite (Cu 2 S: ⬎79% Cu) Bornite; atacamite, azurite Kenhsuite (Hg 3 S 2 Cl 2: 82% Hg); montroydite; metacinnabar Ultramafic (serpentine anomalies) millerite (NiS) Galena (PbS: ⬎73% Pb)

Sphalerite [(Zn, Fe)S: 67.1% Zn]; wurtzite

Automobile exhausts, spoil heaps and tailings, fossil fuel combustion Spoil heaps and mining, pesticides, textiles and electronic waste

Sources: a Swaine (129); b Allaway (130); c Bohn et al. (131).

such high concentrations that they can constitute 1% of their leaf dry weight (12) and are described as hyperaccumulators. The majority of hyperaccumulators consist of small herbaceous plants, although the tree Serbertia acuminata (se`ve bleue) is notable for its bright blue sap, containing approximately 11% nickel as the citrate complex. Elements such as Cu and Zn are required by biological systems for normal function; in plants, these concentrations are low—approximately 6 and 20 mg kg ⫺1 DW, respectively—and are rarely deficient. Copper is required for several enzymes involved in oxidation and reduction reactions, e.g., cytochrome oxidase (13), whereas zinc is required in greater quantities, with more than 80 plant enzymes containing this metal (14). The anthropogenic sources of heavy metals (Table 1) can be broadly grouped into metalliferous mining and smelting, industry, atmospheric deposition, agriculture, and waste disposal on land (11).

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Degradation of soil quality affects forestry efforts as well as natural woodland ecosystems. Contamination of the soil with heavy metals, radionuclides, and organic compounds is widespread. The United States alone has a total of 1283 soil contamination sites on the National Priority List of the Superfund program (15). These sites include chemical plants, contamination from urbanization, groundwater contamination, industrial waste treatment facilities, landfill, sites contaminated as a result of military operations, radioactive operations, and mine tailings. The extent, diversity, and complexity of the contamination is enormous. When comparing the quantity of heavy metals released to the environment as a result of anthropogenic activities to the natural background, Campbell et al. (16) showed that anthropogenic input was 15 times higher than natural for Cd, 100 times higher for Pb, 13 times higher for Cu, and 21 times higher for Zn. One factor that highlights heavy metals as particularly important contaminants is their immutable nature, i.e., they cannot be degraded and in some cases their complexes are just as toxic as the free metal ion. In the case of Hg, for example, organic complexes are considered to be more toxic than inorganic forms, as demonstrated by Picea abies exposed to HgCl 2 and methyl-HgCl 2 in nutrient solution (17). In this case, heightened toxicity was shown to be a result of the chemical form of Hg rather than an increase in uptake. Heavy metals also have a considerably longer residence time within various components of the biosphere than organic contaminants. In situations where heavy metal deposition has decreased through stricter pollution controls, their presence in forested ecosystems remains a problem. The environmental toxicity of heavy metals is governed by their bioavailability. Put simply, the bioavailable heavy metal concentration is that which is readily accessible for uptake by any biological organism. In the soil, plant roots tend to take up heavy metals present in soluble forms. If the metals are insoluble they can remain bound to soil colloids without entering the soil solution and can be considered in the meantime to be relatively harmless, e.g., the complexed forms CdCl 2 and Zn-EDTA are more bioavailable due to their increased solubility. Heavy metals remaining in a nonavailable form within the soil profile may be solubilized by a variety of processes, the most prominent being changes in soil pH. Generally, the more acidic the soil matrix becomes the more bioavailable the heavy metals become. Hence, the widespread incidence of acid precipitation in many European and North American forest systems becomes a major factor in exacerbating heavy metal contamination. Studies of catchment areas in Sudbury, Ontario have shown that the effects of mining and resource recovery activities, now ceased, have left a legacy of elevated Cu and Ni concentrations in the soil expected to persist for the next 1000 years (18). There is a great deal of experimental evidence confirming that acidic precipitation increases the concentration of bioavailable metals; Roemkens and Salomons (19) demonstrated that in forest soils, Cd and Zn concentrations increased at a pH below 5.5. Multiple linear regression of their data showed that

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the change in cation exchange capacity (CEC) and pH explained 79% and 83% of the measured distribution coefficients (K d ) for Cd and Zn, respectively. Soils with an inherently low acid buffering capacity, such as the highly weathered ultisols of the southeastern Atlantic coastal plain, are far more susceptible to increases in metal bioavailability under conditions of acidic precipitation. Cu in particular binds very strongly to organic material, and can remain in the soil and continue to affect ecosystems many years after the contamination episode (20).

4 4.1

EFFECT OF HEAVY METALS ON WOODY PLANTS Physiological Effects

A great deal of research has been dedicated to the study of how metals affect woody plants, ranging from their effect on specific cellular processes to their effect on nutrient cycling within forest ecosystems. Heavy metals impact woody plants via two main pathways; through the root system from metals entering the soil, and via aerial deposition. By far the most direct effect is produced through aerial contamination (21), whereby metals are deposited onto the surface of the leaves. Aerial deposition of heavy metals also impacts the soil because deposited metals are washed down the trunk and accumulate around the base of trees (22). Heavy metals entering the soil can go through a wide variety of species changes as a result of the prevailing edaphic conditions controlling bioavailability (such as pH, CEC, and organic matter content) (23). Interaction with soil microbes and reactions with root exudates also change the species and bioavailability of heavy metals (24). Once metals reach root tissues, their uptake is dependent on certain physiological factors, including characteristics of the heavy metal in particular. Metals that also function as essential micronutrients, such as Cu and Zn, have established mechanisms of uptake, and increasing external concentrations frequently result in increased uptake; for example, the roots of pine seedlings growing on soils spiked with 200 µM ionic Zn accumulated up to 73 µM g ⫺1 of Zn (0.48% dry weight) (25). Typical metal concentrations of metals within soil solution and plants are shown in Table 2. Nonessential metals, such as Cd, can enter plant cells via nonmetabolic pathways (passive diffusion as well as active uptake); work on passive Cd uptake in excised barley roots described a ‘‘saturation’’ effect where Cd floods into the root cells unhindered above specific concentrations (26). This may result from cell membrane damage caused by the high ambient metal concentrations, which disrupts the selective uptake mechanisms, allowing an uncontrolled efflux (27). It seems likely that mixtures of nonmetabolic and metabolic mechanisms are responsible for Cd uptake. Solution culture experiments comparing Cd uptake by Fagus sylvatica L. (common beech) showed high uptake at 20°C, although

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TABLE 2 Typical Concentration of Heavy Metals in Uncontaminated Soil Solutions and Plants Normal conc. (µg kg ⫺1 DW) Element

Soil (total) a

Plants b

Toxic conc. (soil solution; mg L ⫺1) c

Cadmium (Cd) Chromium (Cr) Copper (Cu) Mercury (Hg) Nickel (Ni) Lead (Pb) Zinc (Zn)

0.01–7 5–1000 2–100 0.02–0.2 10–1000 2–200 10–300

0.2–0.8 0.02–15 4–15 0.005–0.5 0.03–5 0.1–10 8–400

0.001 0.001 0.04–0.3 0.001 0.05 0.001 ⬍0.005

Sources: a Swaine (129); b Lindsay (132); c Allaway (130).

Cd efflux was still detected after the temperature was reduced to 0°C—when most metabolic processes had ceased (27). Punshon and Dickinson (28) found that Cd uptake in Salix trees occurred during exposure to an incremental supply of Cd (0.15 mg L ⫺1 increasing to 1.5 mg L ⫺1 over 128 days), although in all tissues the Cd concentration remained consistent at approximately 100 µg g ⫺1. This indicated that—for all Salix species investigated in the study—Cd influx may be restricted by a mechanism that did not place demands on metabolism (i.e., passive) because both leaf and shoot biomass were unaffected by elevated concentrations of Cd when compared with control-grown plants. In the same experiment, Cu accumulated in the root tissues (⬇300 µg g ⫺1) and Zn accumulated in leaves, stems, and roots. In Salix burjatica Nazarov. cv. Aquatica leaf Zn concentrations greater than 600 µg g ⫺1 were accompanied by severe signs of toxicity. Once inside the symplasm, heavy metals inhibit growth and reproduction via a variety of mechanisms. Root elongation is frequently used as a first indication of a plant response to soil-borne metals because this is the organ that first comes into contact with soil contamination. Metal resistance characteristics of plants were initially based on differential root elongation indices in metal-treated and untreated solutions (29). The exact mechanisms of growth inhibition are not as yet fully described (30), but it is likely that mechanisms are specific to the metal(s) and the plant species involved, and also as a function of the external metal concentration. In general, heavy metals seriously disrupt a variety of physiological processes because they bind very strongly to enzymes. Van Assche and Clijsters (31) summarized the effects of metals on enzyme systems, reporting

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that they bind to functional ligands (mainly sulfhydryl groups) that regulate structure and catalytic action, also inducing deficiencies of other metal ions by substitution into metalloproteins. Cross-linkage enzymes responsible for cellular elongation (32) may be also affected in this way by metals, subsequently inhibiting the elasticity of developing root cells. The loss of elasticity as a result of metal toxicity has been demonstrated in the cells of herbaceous plant roots, specifically Phaseolus vulgaris (33) and Festuca rubra (34). Metal contamination of the soil may also inhibit the overall development of root biomass, as reported in Picea abies in response to Cd and Zn (35) and in a variety of other forest species (e.g., Pinus strobus, Pinus taeda, Liriodendron tulipifera, Prunus virginiana and Betula alleghaniensis) in response to Cd (36). Often the reductions in root biomass are severe and occur at relatively low bioavailable concentrations. Wickliff and Evans (37) reported that in nitrogen-free solutions, as little as 0.03 ppm Cd resulted in a 44% reduction in root biomass and accumulation of up to 61 mg kg ⫺1 Cd within dried root tissues. Other effects of heavy metals include a change in root architecture, inhibition of root hair formation, and reduced root initiation (30). These symptoms are not universal to all species; adventitious roots of Salix spp. cuttings respond to exposure to Cu, Cd, and Zn in solution by significantly increasing the number of roots (28,38,39). Heavy metals can also reduce the percentage of mitotic cells in the root apical meristems (40), a reduction in root hair formation, root membrane damage (41), and decreased transpiration rate (42). Therefore, the reduction in root biomass is the overall effect of several individual physiological effects caused by heavy metals. Populus spp. trees exposed to elevated concentrations of Ni (5 mg L ⫺1) in nutrient solution have a drastically reduced transpiration rate in comparison with control plants—with control tree cuttings transpiring approximately 0.5–1 L plant ⫺1 day ⫺1, and Ni-treated plants only used several milliliters per day (Punshon and Adriano, unpubl.). Metals also affect photosynthesis by inhibiting enzymes, producing visible symptoms of toxicity such as stunting, chlorosis, and necrosis (31). 4.2

Effects on Forest Ecosystems

In the late 1970s, evidence of widespread environmental damage began to appear in the coniferous forests of northern and northwestern Europe. This took the form of visible toxicity symptoms such as yellowing of the needles and advancing tree mortality, causing a large-scale decline in the health of conifer plantations. First observed in spruce forests of Bavaria, West Germany, forest decline has spread across Europe and into the United States (43). By the early 1980s, 20–25% of European forests were classified as moderately or severely damaged from seemingly unknown causes. Research indicated that forest decline may be caused primarily by air pollution—specifically SO 2 and NO x , although correlations between

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tree mortality and any one specific stress factor could not be found (44). Schulze (43) believes that forest decline is a response to a complex set of physiological effects triggered by ozone, SO 2 , NO x , ammonia, and organic pesticides. Until recently, heavy metals were not implicated as a contributory factor in forest decline, although workers have now found elevated concentrations of phytochelatins (produced during metal stress to chelate metals within the cell) in the needles of red spruce collected in the United States, suggesting that heavy metals may be partly responsible (45). There is a wealth of information in the scientific literature about the dynamics of heavy metal movement and accumulation in forest ecosystems. Studies tend to concentrate around northern Europe and the former USSR, where acid rain has amplified the effects of metal deposition (19), although studies have spread to the United States where the effects of decline are beginning to emerge. Many studies focus on the effect of aerial deposition from nearby smelting plants on existing forested ecosystems (18,20,46–48). One effect common to many studies is the effect of heavy metals on decomposition processes. Nutrients cycle through the forest ecosystems via breakdown of leaf litter, representing a major process for returning nutrients to the soil. The recovery of nutrients from senescing leaves can contribute 60–80% of nutrients required by the following season (49,50). In metal-contaminated situations the rate of litter decomposition drops dramatically and nutrient cycling is diminished. Often this is the result of direct toxicity of heavy metals to the detritivores that feed on leaf litter and convert it into humic material (51), and decomposition is reduced to such an extent that large quantities of litter accumulate on the surface. Nikliska et al. (52) found that treatment of forest litter with Cd, Cu, Pb, or Zn significantly reduced the respiration rates (used as a measure of degradation metabolism). Respiration rates were even lower after the contaminated litter was stored for progressively longer periods of time. There is evidence to suggest that a reduction in decomposition may be a common response of many different forest ecosystems and that metals tend to be trapped within the system because nutrient cycling slows down; this characteristic ‘‘trapping’’ of metals in the decomposer system has been shown in the mangrove swamps of southeastern Brazil. Silva et al. (53) studied the cycling of Zn, Mn, and Fe through the production, decomposition, and export of litter in the Itacurussa Experimental Mangrove Forest and found that less than 0.01% of metals were exported from the system via the leaf litter degradation pathways. They concluded that this represented a biogeochemical barrier to metal transfer. However, it is likely that in aerobic soils metal availability and hence metal export may be much higher, and the retention of bioavailable metals within the system will have a deleterious effect. Heavy metals also affect the microbial communities in forest soils. Transection studies along pollution gradients have shown that the community structure and the tolerances of bacterial species change on a distance gradient from a metal

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smelter (54). These studies showed that bacterial species were less sensitive to heavy metals than fungal species; that with increasing proximity to the metal smelter, fungal phospholipid fatty acid content of the soil (used as an indication of the magnitude of fungal biomass) was significantly reduced; and that bacteria in this region of the gradient exhibited increased metal resistance in comparison with populations farther away. Since fungal species in the soil are important for mycorrhizal colonization of tree roots and in decomposition of leaf litter, there are severe repercussions to increased heavy metal loading of forest systems. 5

ADAPTATIONS OF WOODY PLANTS TO HEAVY METALS

Heavy metal tolerance of plants as a result of anthropogenic pollution has been studied since the 1930s when stable ecotypic adaptation was demonstrated in the herbaceous plant Silene dioica growing on a metalliferous mine site (55). It is thought that through a process of natural selection certain plants can take advantage of metal-contaminated niches and retain tolerance traits within their population, forming an ecotype. Laboratory studies have shown stability of tolerance traits in herbaceous plants, by exposing tolerant and normal ecotypes of the same species to metals in nutrient culture and observing their differential growth (3,56). Normal populations of plant species contain individuals that express specific genetic mutations dependent on the extent of genetic variability. Certain mutations allow individuals to survive elevated concentrations of heavy metals in the soil, although they may only number one or two for every thousand individuals within a population (57). Although seemingly insignificant, these individuals can initiate colonization of contaminated niches. For trees, however, there is little evidence that they can develop tolerance in the classic sense, i.e., tolerance that is passed on to progeny (30). An important example of this was reported by Watmough and Dickinson (58), where chronic aerial Cu contamination from a nearby smelter impacted a mature sycamore (Acer pseduoplatanus L.) woodland, resulting in the induction of Cu resistance traits in established trees but effectively halting any seedling regeneration. In this case, certain adaptive mechanisms were employed by the mature trees, which were not passed on to their seed, indicating that ‘‘true’’ (heritable) tolerance was not responsible for their continued survival. Instances of metal tolerance via compartmentalization of Zn as an organic complex within the cell vacuole have been reported for barley (Hordeum vulgare L.) (59) and other grasses (60), whereas reported instances of woody plant resistance to metals were mediated through symbiotic mycorrhizal association, shown primarily in Betula (61–63). Jones and Hutchinson (64) also worked with birch (B. papyrifera Marsh.) inoculated with the mycorrhizal fungus Scleroderma flavidium E. & E., known to increase plant resistance to Ni. They found that inoculated plants were more metal-resistant than controls, not because the fungus al-

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tered the uptake and distribution of Ni within the plant as was previously thought but because it enhanced uptake of P and Mg within root tissues. The importance of mycorrhizal association cannot be overstated when discussing the relationship between woody plants and heavy metal contamination. Remediation schemes involving trees are dependent on the use of mycorrhizal incoluation for their success (65). Some species of trees naturally invade industrially contaminated or disturbed land, and several workers have found Betula and Salix species growing of mine spoils enriched with heavy metals such as Pb and Zn (66) as well as Cd (67). There are examples of tree species with much higher metal resistance capabilities than others, many of which involve Betula or Salix species. In nonmycorrhizal studies involving a variety of willow species, Punshon and Dickinson (68) found wide variation in responses, consistently higher metal resistance in a mine ecotype of Salix caprea, taken from a Pb-Zn mining site (Trelogan, Clywd, UK). This ecotype was able to withstand plant-available concentrations of Cu, Cd, and Zn in solution culture in which nonadapted species could not survive; root elongation–based tolerance indices in excess of 100% for the ecotype indicated that growth in metal-enriched solutions was greater than in controls. The mine spoil substrate in which the ecotype had been growing contained total concentrations (µg g ⫺1 DW) of 72.2 (⫾40) Cu, 116.6 (⫾7.1) Cd, 1468 (⫾88) Zn, in addition to 13,488 (⫾527) Pb, and only contained 4% organic matter (39). Further analysis of the roots and the presence of fruiting bodies indicated that they were associated with a mycorrhizal fungus, thought to be Hebeloma spp., indicating that even given elevated metal resistance, trees still require mycorrhizal association, especially when growing in contaminated soil. Physiological studies of trees have revealed that they respond developmentally to changes in their environment. For example, they respond to variation in light incumbent on the canopy by changing the shape of their leaves and even the concentration of chloroplasts within the leaves. These sun-and-shade leaves have alterations in the surface area to either maximize photosynthetic area (required in the shade) or reduce it (required during long hours of sun exposure) (69). A similar response in leaf morphology is seen in eastern cottonwood when exposed to different soil moisture regimes. In this case, leaf size increased from xeric provenances in the west to the mesic provenances in the east (70). These morphological adaptations can be seen within an individual; branches represent independent meristems arising from the central trunk and respond independently to stresses presented by their immediate microclimate. This is known as the theory of genetic mosaicism (71). Such predisposition of woody plants to adaptation within the lifetime of the individual may be a product of their longevity and/or a way of avoiding the metabolic costs of heritable tolerance (72). Considering this, it seems feasible to assume that if a tree species possesses the appropriate genetic variability it can potentially adapt to environmental stress, such as heavy metal contamination, and that this adaptation can be manipulated.

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Several studies have shown that trees from the genus Salix (including Populus) are highly variable, have a predisposition for poor and contaminated soils, and can be acclimated to withstand increasing metal concentrations over relatively short periods of time (28). The occurrence of Salix spp. on industrially contaminated sites (73) is evidence that they possess the genetic variability to invade these niches. Presuming that long-lived woody plants need to adapt to stress within their lifetime, other mechanisms of tolerance, such as phenotypic plasticity, have been considered as an alternative mechanism. Phenotypic plasticity is the variable expression of the phenotype in response to stress and allows rapid adaptation without being passed on to progeny. In a review, Thompson (74) implicates phenotypic plasticity as a possible component in the development of metal resistance in woody plants (75). 6

BENEFICIAL USES OF TREES IN SOIL REMEDIATION

The various environmental benefits of the establishment of woody vegetation has already been mentioned, with emphasis on the fact that trees perform a variety of vital functions in the maintenance of ecosystem health. It is the application of these functions to contaminated soils that has given rise to the use of trees as agents of environmental clean-up. There are two main innovations that use plants to remediate metal contaminated soils, specifically involving fast-growing trees such as Salix (willows and osiers) and Populus (poplar), but which also use Alnus (Alder) and in some cases Eucalyptus (Eucalypts). They are short-rotation forestry (SRF) as a means of disposing of sewage sludge and phytoremediation. SRF was initially used as a means of producing CO 2-conservative energy but later developed into a means of sewage sludge disposal when the high nutrient and water requirement became costly. The heavy metal resistance of the trees used in SRF then became significant due to the extremely high heavy metal concentration of domestic and industrial sludges. SRF was pioneered by the International Energy Agency, an initiation involving Austria, Canada, Denmark, Finland, Italy, the Netherlands, Sweden, United Kingdom, United States, and Brazil (76). This emerged from the rapid development of SRF in Sweden. This resulted from the introduction in 1991 of a new agricultural policy that introduced lower grain prices, compensation for willow plantations on set-aside land, and higher taxes on fossil fuel. Sweden now has 10,000 ha of commercialized energy (or ‘‘biofuel’’) plantations, with the further development of 200,000 ha in progress to replace imported oil traditionally used for heating purposes (77). During 1994 in Sweden, about 80 TWh of energy was generated using biomass fuel, i.e., approximately 20% of the total energy supply (78). However, the energy forestry program was originally established in 1978 with the aim of acting as an advisory body for the dissemination of information on ‘‘high and sustainable production of woody biomass that is

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ecologically acceptable.’’ The large quantities of fertilizer and water required to sustain the intensive management of bioforestry prompted the use of sewage sludge, having the additional benefit of providing a valuable avenue for waste disposal. Information on the use SRF as a means of disposing of wastewaters and sludges was brought together in 1994, prompting wider research on the dynamics of heavy metal uptake and tolerances of various species within the genus Salix (79). At the same time the science of phytoremediation was being developed. The study of heavy metal tolerance indicated that plants could have beneficial practical applications in metal removal from the soil and was coupled with the search for specific hyperaccumulator plants for a variety of toxic metals (80). Phytoremediation initially used herbaceous plants, most notably Brassica juncea (Indian mustard), with a great deal of success (81–85). Trees belonging to the genus Salix have characteristics that recommend them for remediation of inorganic contaminants, especially of Cd, to which they are particularly tolerant (28,86–88). To date, there is still a lack of research targeted at systematically identifying exactly which Salix and Poplar species (and hybrids) are most suitable for use in heavy metal processing and which species could remediate mixed wastes, i.e., organic and inorganic contaminants. Even more important for phytoremediation, information on the use of fast-growing trees in various aspects of environmental remediation, agronomic practices, and their effects on the environment have not been previously collated in a form that custodians of contaminated sites can easily use. Without this knowledge the commercial use of trees to remediate various types of contaminants in a realistic heterogeneous soil environment is still far off. 6.1

Short Rotation Forestry and Waste Disposal

The use of willows, poplars, and other fast-growing trees in the disposal of wastewaters and sludges is based on the practice of SRF, which has been used successfully in Sweden for the production of biofuel for the past 20 years (89). SRF takes its name from the short rotation time typical of poplar and willows compared with conifers (60–120 years) or broadleaved trees (35–50 years); typically between 3 and 6 years (90). Under optimum conditions (i.e., intensive management strategies), SRF can yield between 12,000 and 20,000 kg dry matter (DM) ha ⫺1 yr ⫺1 (91), although in general, lower yields of between 6000–9000 kg DM ha ⫺1 yr ⫺1 are more likely (90) (Table 3). If SRF is used for the generation of energy, approximately 4.5 MWh of energy can be produced from burning 1 ton DM (50% moisture) (92). There is now a substantial body of data that indicates that the irrigation of biomass plantations with sewage effluent may not only increase and accelerate the growth of the trees but also provides a valuable avenue for recycling an

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TABLE 3 Tree Species Used in Short-Rotation Forestry (SRF) Around the World Country Austria

Salix spp., Populus spp., Alnus spp., Betula spp.

Canada

Populus euramericana (DN17; DN74, DN177, DN181); Populus nigra ⫻ maximowiczii (NM6) S. viminalis S. dasyclados Wimm., S. viminalis, S. myrsinifolia P. deltoides Bartr. ⫻ P. nigra P. nigra, P. deltoides, P. alba Usually P. euramericana, also P. deltoides ⫻ nigra and P. trichocarpa ⫻ deltoides clones Salix, Populus, Alnus, Betula, S. viminalis, S. dasyclados Salix, Populus, Alnus, Eucalyptus, P. deltoides ⫻ nigra, P. trichocarpa ⫻ deltoides, P. ⫻ trichocarpa, P. tachamahaca ⫻ trichocarpa, P. alba ⫻ tremuloides P. deltoides, P. balsamifera, P. trichocarpa, P. tremuloides, Robinia pseudoacacia, Geditsia trianthus, Leucacena refusa, Prosopis glandulosa, Prosopis alba, Betula pendula, Alnus glutinosa, Alnus rubra, Eucalyptus grandis, E. saligna

Denmark Finland Italy Netherlands Sweden United Kingdom

United States

Yield (t DM ha⫺1 yr⫺1)

Tree species (and clone) used

5 without intensive management; 23–24 with; average 10–12 No figures available ⬇7.6 (range: 6.7–8.1) 4.7 to ⬇9–10 3.5–9 8–10 5.6–10 3.6–10

4.5–16.7

Punshon

Note: All yields are based on approximate ranges irrespective of the species of clone used. Source: Handbook on How to Grow Short Rotation Forests, International Energy Agency, 1997 (76).

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abundant waste stream. Production of sewage sludge is approximately 6 million t yr ⫺1 in the United States, 2.5 million t yr ⫺1 in Germany, and 1.07 million t yr ⫺1 in the United Kingdom, and as such its disposal has often been problematic because it requires extensive treatment before it is considered safe. This includes the removal of excess N, heavy metals, organic chemicals such as pesticides, and pathogenic bacteria prior to application onto agricultural land (93). Current research on the use of sludge as a fertilizer for SRF indicates strongly that willow plantations may be capable of performing the majority of the required clean-up steps. Perttu (94) suggested that SRF could be used as ‘‘vegetation filters’’ to utilize the nutrients in municipal sewage sludge, wastewater, leakage water, and bioash (wood ash), although dredged sediments containing organic chemicals and pathogenic bacteria have also been effectively ‘‘purified’’ following application on to biomass plantations. The tree species used in biomass forestry vary between areas in which they are grown. A summary of the species used (Table 3) indicates that in warmer countries, such as the United States, species such as Eucalyptus, which are native to Australia, can be successfully planted. As previously mentioned, the amount of woody biomass produced in SRF varies with species, management practices, and location, although maximum yield values of 23–24 t DM ha ⫺1 yr ⫺1 have been noted in Austria (76) (Table 3). Many of the SRF species commonly used, especially Populus spp., are those that have also shown considerable potential in the remediation of groundwater contaminated with organic chemicals, specifically the P. deltoides ⫻ nigra and P. trichocarpa ⫻ deltoides clones (95). SRF fulfills several important environmental and ecological factors; there is no net CO 2 contribution to the atmosphere compared to energy production using fossil fuel, and there is a small net uptake. Pesticide use is lower than on conventional agricultural crops, and the growth of willows for SRF improves the condition of the soil as well as biodiversity (96,97). Primary requirements of SRF are a rich supply of nitrogen and water. The ability of fast-growing trees to remove nutrients and water can also be applied to wastewater treatment; excessive concentrations of nutrient chemicals pose environmental problems when present in excess, such as eutrophication of streams and lakes, causing algal blooms and a progressive reduction in potability. Standards usually monitor the biological oxygen demand (BOD), total N, ammonia N, and total P in water (98), and often these strict regulatory limits cannot be met by conventional treatment methods alone. The elimination of P from wastewater generally involves the use of large quantities of chemical agents, which are difficult to clean up once the process has been completed. Obarska-Pempkowiak (99) found that Salix viminalis and S. arenaria could significantly reduce the nitrogen concentration of municipal wastewater; in particular, S. viminalis reduced the total N concentration of wastewater from more than 35 gm ⫺2 to less than 5 g m ⫺2 over a period of 270 days. Typical N and P contents of applied sludges are

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approximately 4.5 kg N t ⫺1 and 2.8 kg P t ⫺1 in the raw cake, of which 1.5 and 1.4 kg t ⫺1 N and P, respectively, are available in the first cropping season (100). Kutera and Soroko (98) found that willow and poplars irrigated with sewage sludge showed an increased yield and that the trees took up almost 95% of the N and P load of the effluent, simultaneously lowering the BOD of the effluent by an average of 96%. Removal of water by fast-growing trees remains an important issue, especially for those wishing to use the trees to obtain some degree of hydrological control, e.g., in contaminated groundwater seepline interception. This involves the use of fast-growing phreatophytes to utilize contaminated groundwater, degrade the organic chemicals, and dry out the seepline area, where there is often standing water. Exact figures of water use vary, and depend very much on the soil conditions, especially the presence of other heavy metals contaminants. SRF studies use water use efficiency (WUE) as a means of measuring transpiration rates, measuring the amount of dry biomass produced per unit of water. Studies on the average long-term WUE of a stand of Salix viminalis showed a production of 6.3 g dry biomass per kg of transpired water, which is high compared with other tree species (101). However, for hydrological control and seepline interception, species choice criteria obviously include a tree with a lower WUE, i.e., the highest throughput of water per gram of dry biomass. Information about nitrogen status is important when considering water use because foliar nitrogen concentration affects stomatal conductance (101). Work by Blake et al. (102) showed that WUE varies widely between clones within a particular species and demonstrated this for Populus. They found that the cottonwood P. nigra N-80 (section Aegeiros) had a low WUE, in that biomass production was lower per kg transpired water, whereas the balsam poplar P. maximowiczii M-4 (section Tacamahaca) and the white poplar P. alba A-499 (section Leuce) had twice the dry matter production for the same amount of transpired water. 6.2

Removal of Heavy Metals

Ecological studies have identified fast-growing trees invading metalliferous mine spoils (67); therefore, the scope of their use in waste disposal and soil amendment extends also to their removal of metals. One of the most notable features of Salix is their seemingly innate ability to take up concentration of Cd within their tissues far in excess of the range considered normal. Heavy metals taken up during biological purification of sewage sludge were investigated by Nielsen (103) who found that S. viminalis accumulated significant concentrations of Cd within root and shoot tissues. Studies have emphasized that SRF can be used for ‘‘collecting’’ metals from the soil, in effect using the trees as biological filters. Go¨ransson and Philippot (104) found that almost all of the Cd applied during sludge application was taken up by Betula pendula and that theoretically the trees could remove

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1.5 kg Cd ha ⫺1 yr ⫺1 when watered with sludge. Landberg and Greger (105) found that net transport of heavy metals to the shoots varied widely between 1% and 72%, therefore, certain clones were capable of accumulating more metals than others, indicating the need for clone selection for resistant trees (106). Systematic screening of fast-growing trees to heavy metals is lacking; discovery of species able to tolerate and remove metals remain serendipitous, and responses of Salix to metals suggest a species-specific response. In hydroponic experiments, Salix cuttings were exposed to varying levels of Ni for 28 days with measurements taken of the length of the longest root and shoot. Responses of these important SRF clones to Ni supplied at external concentrations of between 0–5 mg L⫺1 show that above 1 mg L⫺1 growth inhibition is severe (Figure 1). Maintenance of sub-lethal external metal concentration is a pre-requisite in the use of nonhyperaccumulator species, especially where harvesting the woody biomass is used as a means of metal removal. External metals concentrations affect the rate

FIG. 1. Effect of varying concentrations of nickel on leaf and root growth in Salix, grown hydroponically and harvested after 28 days. Data are means of longest root and shoot ⫾ SE where n ⫽ 5.

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of uptake, with lower concentrations allowing a greater total accumulation. Resistance induction studies using Cu, Cd and Zn have shown that if low initial external concentrations of metals are used, and doses are gradually increased over time, growth responses can be considerably enhanced in comparison with sudden exposure. By increasing concentrations ten-fold over a 128 day period, the SRF clone Salix viminalis showed no signs of root growth inhibition (Figure 2) in response to heavy metals (39). There is very little information on metal uptake characteristics of Alnus (37,107) or Eucalyptus (108,109). To date, these species have not been screened for their ability to remove heavy metals from the soil. The efficiency with which willows can remove metals from sewage sludge fertilized soils has been demonstrated in several key studies. Table 4 summarizes two studies on willow vegetation filters to give an indication of the metal concentration contributed to the soil by application of sewage sludge and how much willows can remove. In a field experiment using S. viminalis, S. viminalis ⫻ triandra, and S. dasyclados, Riddel-Black (106) found that the trees were able to remove 2.7–3.3% of the total Ni added via sludge application; 3.2–4.5% of the Cu, up to 45% of the Zn and most notably a maximum of 426% of the Cd added via sludge was removed by S. viminalis ⫻ triandra Q83. Table 4 also confirms this affinity for Cd, which indicates that of the 5.2 g Cd ha ⫺1 added

FIG. 2. Effect of gradually increasing nutrient solution concentrations of heavy metals on root length in the biomass willow Salix viminalis. Data are mean longest root length ⫾ SE where n ⫽ 54. Initial and final metal doses are indicated.

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TABLE 4 Comparison of Concentrations of Heavy Metals Added by Sewage Sludge Fertilization and Removed by SRF a Heavy metals Source 1. Nielsen [103] Using S. viminalis (78–112) Added via sludge (6,833 kg ha⫺1 DW) (g ha⫺1) Removed with biomass (g ha⫺1) 1. Riddel-Black (106) Using S. viminalis and S. triandra hybrids (Bowles, SQ 683, Triandra Q83) Added with sludge (200 m3 ha⫺1) (g ha⫺1) Removed with biomass (g ha⫺1): viminalis ‘‘Bowles Hybrid’’ viminalis ‘‘SQ 683’’ viminalis ⫻ triandra ‘‘Q83’’ dasyclados a

Cu

Cd

Pb

535

5.2

447

4.6–13.1

2.0–6.0

16.2–47.7

140

58.6 63.1 81.5 73.6

6

53 34.3 76.7 6.01

Ni

93

13.4 9.3 11.1 13.8

Zn

419

1.56 0.954 1.36 1.21

Values are expressed as g ha⫺1 except where stated otherwise.

(103), up to 13.1 g Cd ha ⫺1 was removed. Therefore, it was calculated that the number of years of willow SRF practices required to reduce heavy metals loads of the soil to the target concentration is far shorter for Cd than for the other metals (106). There is evidence to suggest that fast-growing trees may be capable of playing a major role in reducing the concentration of organic chemicals in contaminated sludges, soils, and groundwater (see next section). Using landfarming techniques, Harmsen et al. (110) used willows to remove polyaromatic hydrocarbons from dredged harbor sediments, concluding that the high yield of biomass was instrumental to reducing the costs. Willows and poplar trees were also used in a multispecies wetland system to remove bacteria and other pathogens from municipal wastewater. 6.3

Fast-Growing Trees in Phytoremediation

Phytoremediation is a relatively new technology that involves the use of green plants to ameliorate the condition of the soil. Usually the amelioration takes the form of degradation or removal of contaminants, although it can also be extended to include improvement of the physical condition of the soil or immobilization of contaminants. The study of phytoremediation is generally divided in three

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distinct areas: phytoextraction, phytodegradation, and phytostabilization. Phytoextraction involves the removal of contaminants from the soil solution and in to the tissues of the plant and is generally associated with inorganic contaminants, such as heavy metals. Phytodegradation techniques are usually applied to organic contaminants, where aspects of the plant metabolism are used to convert contaminants to benign byproducts. Phytostabilization is the use of the physical presence of plants to improve soil quality and can entail erosion prevention in situations where contaminated material is being spread via the action of wind or water. Heavy metal–contaminated sediments are usually accessible to either phytoextraction (to remove metals) or phytostabilization (to immobilize the contaminated soil or the contaminants within it). Studies of phytoremediation also often include the term rhizofiltration to describe the use of plants to remove heavy metals from polluted waters. One of the most attractive features of phytoremediation is the incredibly low cost in comparison with conventional clean-up technologies. Hazardouswaste clean-up is projected to cost at least $400 billion in the United States alone; $7.1 billion on sites contaminated solely with heavy metals and $35.4 billion on sites contaminated by metals and various types of organic contaminants (111). In addition, the environmental impact of some of the conventional technologies means that the soil is eroded, degraded, or biologically inert once clean-up has been completed, and therefore it still requires revegetation and stabilization in addition. In a review of phytoremediation of heavy metals, Salt et al. (111) state that ‘‘the optimum plant for the phytoextraction process should not only be able to tolerate and accumulate high levels of heavy metals in its harvestable parts but also have a rapid growth rate and the potential to produce high biomass in the field.’’ Initially, however, hyperaccumulator plants and plants from the family Brassicaceae were the main species to be used in phytoextraction field trials. This choice was prompted by solution culture studies in which the hyperaccumulator Thlaspi caerulescans accumulated 2739 (⫾ 323) µg g ⫺1 DW Ni after 8 days of growth in a solution containing 1 mg Ni L ⫺1 (111), although the biomass production of the plants was very low. Fast growing trees, such as willows (e.g. Salix caprea) grown in Ni-enriched hydroponic solution, took up less Ni over a longer period of time than Thlaspi; ⬇ 120 µg g⫺1 DW in the leaves of plants grown in solution containing 4 mg Ni L⫺1 for a period of 21 days (88) (Fig. 3). However, despite this, willows produce several orders of magnitude more biomass than the hyperaccumulators, and therefore metal removal using these plants may still be feasible. Unadapted nonmycorrhizal willows accumulated Ni above the normal range of 0.02–5 µg g⫺1 in hydroponics (11). From these studies it would appear that S. caprea could accumulate reasonably high concentrations of Ni provided the solution concentrations was not toxic; accumulation was still approximately 20 times higher than background at this treatment level. Heavy metal uptake

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using nonadapted plants is achievable as long as excess toxicity is avoided. Nickel availability in the soil varies with the prevailing edaphic conditions but is generally several orders of magnitude below the total concentrations (112), and given a plentiful supply of water and available N, fast-growing trees may theoretically be useful for phytoextraction in situations of low to moderate metal contamination, while producing enough woody biomass to make the process economically feasible. Other phytoremediation studies have indicated that hybrid poplars may be useful for heavy metal clean-up; Banuelos et al. (113) tested a range of hybrids in culture using saline-, boron-, and selenium-enriched drainage water and found that P. trichocarpa ⫻ deltoides performed best in terms of the lack of toxicity symptoms. Hybrid poplars (members of the genus Salix) have been widely studied for their ability to degrade organic contaminants, most notably trichloroethylene (TCE) (95,114) but also tetrachloroethylene (PCE) (115), trinitrotoluene (TNT) (116,117) atrazine (118), and 1,4-dioxane (119). The phytodegradation of volatile organic compounds (VOCs) is not likely to be due to the possession of a unique metabolic pathway able—such as in the case of TCE—to perform reductive dechlorination, but rather a function of rapid growth, a deep root system that can rely entirely on groundwater aquifers (i.e., they are phreatophytes), and a high transpiration rate in comparison with that of many other tree species. Other species, such as Salix and Eucalyptus, are also phreatophytes and have a characteristic dimorphic root morphology. This refers to the two types of root systems they typically possess: shallow lateral roots that primarily obtain water from precipitation (during the wet season), and the deeply penetrating tap (sinker) roots, which obtain water almost exclusively from the underlying water table (to a maximum of about 30 feet) during drier periods. This innate ability to degrade organic chemicals may provide useful in the remediation of mixtures of organic and inor-

FIG. 3. Metal concentrations (µg g ⫺1 ) in tissues of Salix caprea after 21 days growth in solution culture. Data are means where n ⫽ 3 (88).

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ganic contaminants. However, there is a shortfall of research information on how this may be achieved and which clones could be used; organic degradation relies on high transpiration rates, whereas many heavy metal contaminants severely reduce water uptake. It also indicates that if phytodegradation schemes are to be successful, they can only be carried out on sites without toxic levels of heavy metal contamination. Eucalyptus has been traditionally used to dry out water-logged soils because of its extremely high transpiration rates, although only a relatively small number of studies have been carried out using Eucalypts for soil remediation. One specific study utilized their high transpiration rates as conduits for mercury transpiration (120), and another studied Eucalypts for reclamation of mine spoils (121–124) with good rates of survival and biomass production. Mishra et al. (125) used hybrid Eucalypts in reclaiming coal fly ash and achieved successful long-term site revegetation. However, there is little additional evidence to suggest that Eucalypts have an ability to tolerate elevated concentrations of inorganic or organic contamination and still maintain their high transpiration rates. Abouelkhair (126) found that Cd concentrations up to 16 mg kg ⫺1 decreased seedling height of Eucalyptus camaldulensis inoculated with the ectomycorrhizal fungus Pisolithus tinctorus, although it was also noted that stem diameter, weight, leaf area, and water uptake were not affected. Furthermore, Mitchell et al. (127) used Eucalyptus eximia as a sensitive species in assessment of phytotoxic copper, triallate, and anthracene concentrations. It is therefore unlikely that Eucalypts— although ideal for biomass and hydrological control—can be used in remediation programs where heavy metals are a concern. 7

CONCLUSION

It is clear that the extent to which fast-growing tree crops can be used for phytoremediation of heavy metals has not been fully explored. Our ecological knowledge of these plants indicates that certain species possess useful metal resistance characteristics and can rapidly establish on degraded contaminated land, ameliorating the contamination over successive cropping periods and stabilizing the soil. Their innate ability to metabolize organic contaminants and tolerate elevated concentrations of heavy metals, such as Cd, should stimulate research into mixed waste remediation, as it is the complex heterogeneous substrates that provide the greatest challenge for phytoremediation. By far the largest portion of the projected cost in conventional environmental clean-up ($35.4 billion) is taken up by these sites. The potential for the production of biomass during phytoremediation schemes may become an important selling point for a technology for which enduser acceptance remains a problem. The information reviewed in this chapter is offered as an attempt to introduce further considerations in the use of trees for

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environmental clean-up, such as water use, nutrient requirements, and other potential applications, because it is this information that will ensure the success of clean-up schemes using fast-growing trees. The presence of these two very similar fields of research developing concurrently but independently suggests that more rapid developments may be possible if the economical, ecological, and agricultural lessons learned from SRF are used as a starting point for phytoremediation. There is some evidence to suggest that the initial signs of crossover are occurring; work using ‘‘land-farming’’ techniques (biomass production for energy using willow crops) have been suggested for the removal of polyaromatic hydrocarbons and mineral oils from dredged sediments (110). This represents one of the first applications of acknowledged SRF practices to organic contaminant remediation. There is every indication that biomass production and disposal of both organic and metalliferous wastes is theoretically possible, opening the way for plantations that are specifically established to receive waste material and produce energy from it. The final step in the use of trees to manage our waste products is the safe combustion of the wood for energy and the possibility that inorganic contaminants can be rescued from the ashes and recycled. Researchers in Sweden are already making progress toward the development of safer biomass combustion technology; using a heavy metal separation technique involving high-temperature cyclone separation of metals while retaining Ca, Mg, P, and 75% of K within the ash for reuse (128), recycling valuable nutrients into the intensively managed soil. The ability of fast-growing trees to take up heavy metals from the soil in the presence of organic contaminants remains to be tested in either the laboratory or the field, and it is this interdisciplinary research that will be instrumental in making phytoremediation a more acceptable clean-up technology. The immense variation, hybridization, and adaptability that exist within fast-growing trees could provide economically viable solutions for a wide variety of contamination issues (113). ACKNOWLEDGMENTS The author gratefully acknowledges the assistance of Dr. John Seaman, Dr. Ken McLeod, Dr. Nick Dickinson, and Prof. N. W. Lepp for their helpful comments on the manuscript. REFERENCES 1. E-D Schulze, HA Mooney. Biodiversity and Ecosystem Function. Berlin: SpringerVerlag, 1994, p. 525.

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97. S Ledin. Environmental consequences when growing short rotation forests in Sweden. Biomass Bioenergy 15:49–55, 1998. 98. J Kutera and M Saroko. The use and treatment of wastewater in willow and poplar plantations. In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Sweden, 1994, pp. 37–48. 99. H Obarska-Pempkowiak. Removal of nitrogen and phosphorus from municipal wastewater by willow—a laboratory approach. In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Uppsala, Sweden, 1994, pp. 83–90. 100. D Riddel-Black. Sewage sludge as a fertilizer for short-rotation energy coppice. In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludge: A Biological Purification System, Uppsala, Sweden, 1994, pp. 91– 100. 101. A Lindroth, A Cienciala. Water use efficiency of short rotation Salix viminalis at leaf, tree and stand scales. Tree Physiol 16:257–262, 1995. 102. TJ Blake, TJ Tschaplinski, A Eastham. Stomatal control of water use efficiency in poplar clones and hybrids. Can J Bot 62:1344–1351, 1983. 103. KH Nielsen. Sludge fertilisation of willow plantations. In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Uppsala, Sweden, 1994, pp. 101–112. 104. A Goransson, S Philippot. The use of fast growing trees as ‘‘metal-collectors.’’ In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Uppsala, Sweden, 1994, pp. 129–132. 105. T Landberg and M Greger. Can heavy metal tolerant clones of Salix be used as vegetation filters on heavy metal contaminated land? In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Uppsala, Sweden, 1994, pp. 133–144. 106. D Riddel-Black. Heavy metal uptake by fast growing willow species. In: Proceedings of Willow Vegetation Filters for Municipal Wastewaters and Sludges: A Biological Purification System, Uppsala, Sweden, 1994, pp. 145–152. 107. LH McCormack, KC Steiner. Variation in aluminium tolerance among six genera of trees. For Sci 24:565–568, 1978. 108. J Huang and EP Bachelard. Effects of aluminum on growth and cation uptake in seedlings of Eucalyptus mannifera and Pinus radiata. Plant Soil 149:121–127, 1993. 109. MA Norem, AD Day and KL Ludeke. An evaluation of shrub and tree species used for revegetating copper mine wastes in the south western United States. J Arid Environ 5:299–304, 1982. 110. J Harmsen, A van der Toorn, D Boels, B Vermeulen, W Ma, J van der Waarde, R Duijn, R Kampf. Growing biomass to stimulate bioremediation: technical and economic perspective. In: Proceedings of the Fifth International In situ and OnSite Bioremediation Symposium, San Diego, 1999, pp. 39–44. 111. DE Salt, M Blaylock, Nanda PBA Kumar, V Dushenkov, BD Ensley, I Chet. Phytoremediation: a novel strategy for the removal of toxic metals from the environment using plants. Biotechnology 13:468–474, 1995.

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112. DC Adriano. Trace Elements in the Terrestrial Environment. New York: SpringerVerlag, 1986, pp. 533. 113. GS Banuelos, MC Shannon, H Aiwa, JH Draper, J Jordahl, L Licht. Phytoextraction and accumulation of boron and selenium by poplar (populus) hybrid clones. Int J Phytoremed 1:81–96, 1999. 114. M Gordon, N Choe, J Duffy, G Ekuan, P Heilman, I Muiznieks, L Newman, M Ruszaj, BB Shurtleff, S Strand, J Wilmoth. Phytoremediation of Trichloroethylene with Hybrid Poplars. ACS Symp Series 664:1997. 115. X Wang, LA Newman, MP Gordon, SE Strand. Biodegradation of carbon tetrachloroethylene by poplar trees: results from cell culture and field experiments. Proceedings of Phytoremediation and Innovative Strategies for Specialized Remedial Application: Fifth International In Situ and On-Site Bioremediation Symposium, San Diego, 1999, pp. 133–138. 116. PL Thompson, LA Ramer, JL Schnoor. Uptake and transformation of TNT by hybrid poplar trees. Environ Sci Technol 32:975–980, 1988. 117. PL Thompson, JL Schnoor. Phytoremediation of Munitions (RDX, TNT) Waste by a hybrid poplar. Abstracts of Papers of the American Chemical Society 213: 36-ENVR, 1997. 118. JG Burken, JL Schnoor. Uptake and metabolism of atrazine by poplar trees. Environ Sci Technol 31:1399–1406, 1997. 119. JL Schnoor, EW Aitchison, SL Kelley, PJJ Alvarez, S Wakefield, JG Burken, CL Just. Phytoremediation of 1,4-dioxane by hybrid poplars. Abstracts of Papers of the American Chemical Society 213:195-ENVR, 1997. 120. TL Leonard, GE Taylor, MS Gustin, GCI Fernandez. Mercury and plants in contaminated soils: 1. Uptake, partitioning and emission to the atmosphere. Environ Toxicol Chem 17:2063–2071, 1998. 121. SK Maiti and NC Saxena. Biological reclamation of coalmine spoils without topsoil: an amendment study with domestic raw sewage and grass-legume mixture. Int J Surf Mining Reclam Environ 12:87–90, 1998. 122. SC Ward and JM Kick. Biomass and nutrient distribution in a 15.5 year old forest growing on a rehabilitated bauxite mine. Aust J Ecol 21:309–305, 1996. 123. AJ Williams, RB SIngh, AK Bhowmik, AK Singh, SK Banerjee. Suitability of different tree species for copper mine overburdens. Environ Ecol 12:116–118, 1994. 124. JK Chaturvedi. Afforestation of bauxite mined area in central India. Indian Forester 109:458–476, 1983. 125. PN Mishra, M AK, AJ Williams, D Dugaya, SK Banerjee. Biological reclamation of fly ash by forest plantations. Environ Ecol 13:11–14, 1995. 126. KS Abouelkhair. The influence of cadmium on the growth and mineral content of Eucalyptus camaldulensis seedlings inoculated with Pisolithus tinctorus. Alexandria J Environ Res 33:313–324, 1988. 127. RL Mitchell, MD Burchett, MD Pulkownik, L McCluskey. Effects of environmentally hazardous chemicals on the emergence and early growth of selected Australian plants. Plant Soil 112:195–199, 1988. 128. A Ljung, A Nordin. Theoretical feasibility for ecological biomass ash recirculation: chemical equilibrium behavior or nutrient elements and heavy metals during combustion. Environ Sci Technol 31:2499–2503, 1997.

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129. DJ Swaine. The Trace Element Content of Soils. Bucks, UK: CAB, 1955. 130. WH Allaway. Agronomic controls over the environmental cycling of trace elements. Adv Agron 1968. 131. HL Bohn, BL McNeal, GA O’Connor. Soil Chemistry. New York: Wiley Interscience, 1985. 132. WL Lindsay. Iron oxide solubilization by organic matter and its effect on iron availability. Plant Soil 130:27–34, 1991.

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12 Tree Bark Tree ‘‘Bark Pockets’’ as Pollution Time Capsules for Historical Monitoring

Kenichi Satake National Institute for Environmental Studies, Tsukuba, Japan

1

INTRODUCTION

Historical monitoring of environmental pollution combined with present day monitoring is essential for assessing pollution levels in a world that has different levels and types of human activity and a diverse environmental history. Recently in Japan, bark enclosed in tree trunks, known as tree ‘‘bark pockets,’’ has been shown to provide some of the most readily available historical specimens for monitoring air pollution (1). Bark pockets are common in tree trunks. The phenomenon is well recognized by forestry technicians, saw millers, and dendrologists because trunks with bark pockets often have problems with lack of strength, discoloration, and microbial decomposition of the xylem layer. Bark pockets have not previously been identified as being useful parts of a tree. Their potential value for scientific research has previously been overlooked. Harvested trunks with bark pockets have mostly been discarded as waste. Therefore, many forest trees identified as containing bark pockets have been left standing and not harvested. 353

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This chapter describes the scientific use of bark pockets for environmental research as ‘‘pollution time capsules.’’ This approach opens a new frontier in methods of historical monitoring in the world.

2

TREE ‘‘BARK POCKETS’’ FOR HISTORICAL MONITORING OF AIR POLLUTION

Tree bark provides ideal specimens for directly measuring levels of local pollution. This is because a tree bark accumulates air pollutants directly onto its surface (Fig. 1). However, despite this suitability for providing direct evidence of air pollution, bark has not been used to monitor historical trends in air pollutants because it is also exposed to polluted air emitted by modern sources, thereby confounding historical interpretation. The presence of bark pockets within the trunk removes this disadvantage. During tree growth bark can be incorporated into the trunk to form a bark pocket by many mechanisms that are described in the next section. One of the important aspects of using bark pockets as pollution time capsules is that the bark pockets are located between annual rings, thus allowing extremely accurate dating.

3

MECHANISMS OF BARK POCKET FORMATION

There are many physical and biological causes of the formation of bark pockets as pollution time capsules. The mechanisms are classified as follows: 3.1

Mechanism (1): Bark Pocket—Type 1, Encapsulation of a Wound

Figure 2 shows a wound in a tree trunk caused by physical damage and the subsequent development of covering the wound with newly formed bark and xylem. Figure 3 shows bark pockets formed by this process. Encapsulation of a wound is one of the main mechanisms of bark pocket formation. A wound in the process of encapsulation by surrounding bark is called a ‘‘bark pocket to be.’’ The process of encapsulation of the polluted outer bark leads to the enclosure of pollutants into the trunk by the tree’s annual rings, thus allowing accurate dating. In some cases, biological materials, such as shoots of epiphytic bryophytes, are found between the two bark layers in the bark pocket. Epiphytic bryophytes are potentially indicators of atmospheric pollution (2). Information about the past environment can be obtained from species found in samples through analysis of the pollutants contained in the biological materials. In addition to this, the outermost xylem in the bark pocket (Fig. 3) is available for samples that retain pollutants, provided microbial decomposition is limited.

Tree Bark

FIG. 1 Accumulation of air pollutants on the surface of tree.

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FIG. 2 A wound in a tree trunk caused by physical damage and the subsequent development of covering the wound with newly formed bark and xylem.

3.2

Mechanism (2): Bark Pocket—Type 2, Joining of Two Tree Trunks (Branches)

Figure 4 shows the initial contact between two tree trunks or branches and the radial section of bark pockets formed by this mechanism (X m –X n). Figure 5 shows radial section of bark pocket. Figure 6 shows a vertical section of the bark pocket (Y 0 –Y n). This type of bark pocket provides at least three replicates of continuous time series data on the outer bark samples because the barks are formed symmetrically on both sides of the horizontal and vertical axes. The bark pockets cover long spans of successive years, from old to new, permitting monitoring of continuous, year-by-year, historical changes in pollution. 3.3

Mechanism (3): Bark Pocket—Type 3, Encapsulation of a Branch

Figure 7 show the steps in the encapsulation of a branch. Shaping and/or pruning of branches are the main cause of this type of bark pocket, which is usually observed in trees planted on roadsides and in parks, gardens, and artificial forests.

Tree Bark

FIG. 3 Bark pocket type 1 in a tree trunk.

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FIG. 4 Initial contact between two tree trunks or branches (a) and radial section of bark pockets formed by this mechanism (Xm –Xn) (b).

FIG. 5 Radial section of bark pocket formed by the joining of two branches (x ⫺m –x m).

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FIG. 6 Vertical section of the bark pocket formed by joining two branches (y 0 –y n).

The appearance of the outer bark formed in this encapsulation process differs from that of normal outer barks. This type looks like a ‘‘cat’s eye’’ but it is a ‘‘tree’s eye’’ for watching the environment (Fig. 8). Tree’s eye is an indicator of the location of bark pockets in a tree. Some tree’s eyes in the process of bark pocket formation can be called bark pockets-to-be because they change to bark pockets after encapsulation.

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FIG. 7 Steps in the encapsulation of a branch after pruning.

FIG. 8 Cat’s eye (⫽ tree’s eye ⫽ bark pocket-to-be) formed on the outer bark after pruning the branch.

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The physical and chemical characteristics of outer bark differ markedly with different tree species. For example, the outer bark of black pine (Pinus thunbergii), red pine (P. densiflora) and white birch (Betula tauschii) peel off from the trunk. The barks of some species exfoliate with characteristic periodicity. In this case, it is difficult to estimate trends in air pollution using outer bark and bark pockets because the exposure period of the outer bark is shorter than the time estimated from the annual rings enclosing the bark pocket. However, the section of the cut end of the branch or wax rosin layer in the bark pocket remains available as an accumulator of air pollutants provided that the xylem of the section is not microbially decomposed. In this case, the annual rings of the cut end of the branch and the record of pruning help with correct dating of the section. 3.4

Mechanism (4): Bark Pocket—Type 4, The Joining of Uneven Trunk

Figure 9 shows the bark pocket formed in the process of growing on an uneven trunk. This type of bark pocket is often observed in beech (Fagus japonica, Fagus sylvatica), which has an uneven trunk. This type of bark pocket in the trunk forms in the same compass direction as the continuously jointed outer bark. Bark pockets arranged in the same compass direction in the trunk offer an advantage for understanding historical changes in pollution because there is a directional relationship between the source of pollutants and the tree that accumulates the pollutants.

FIG. 9 Bark pocket formed in the process of growing on an uneven trunk.

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Other Mechanisms (5): Bark Pocket—Type 5

There are other occasional causes and characteristic mechanisms of bark pocket formation. Artificial materials such as wire, nails, and fences can damage a tree and become encapsuled to form bark pockets as the tree grows. Wood fire, thunder attack, and cracks formed in the trunks by freezing during cold winters are other causes of characteristic bark pocket formation. 4

CHARACTERISTICS OF BARK POCKETS AS ‘‘POLLUTION TIME CAPSULES’’

Natural and artificial materials have been used for historical monitoring of pollutants transported into terrestrial and aquatic ecosystems (Fig. 10). Typical of these materials are lake sediments, polar ice, herbarium specimens, peat, tree rings, bones, freshwater shells and corals. However, these materials are not always fit for historical monitoring of air pollution. Difficulties often exist in the use of such specimens for monitoring. Major problems include (a) dispersal and translocation of the pollutants; (b) contamination during natural and/or artificial preservation periods; (c) difficulty in sourcing materials with a suitable time scale from inhabited monitoring sites such as urban areas, and also from remote and background areas; (d) incor-

FIG. 10 Typical natural and artificial materials for historical monitoring of pollutants transported to terrestrial and aquatic ecosystems. (Examples are lake sediments, polar ice, herbarium specimens, peat, corals and tree rings.)

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rect dating. Corresponding examples of these problems are as follows: (a) Human activities, such as fishing by net and bioturbation by bottom fauna, often disturb the laminar structure of lake sediments formed by the deposition of particulates. (b) Herbarium specimens are in contact with the current atmosphere during their storage. Exchange of pollutants between bones and soil or contamination by soil may occur. (c) Polar ice from the Antarctic may provide information on background areas or global pollution, but it is difficult to get information on pollution in populated urban or rural areas. (d) Evaluation and application of dating methods suitable for the selected materials are required. One of the practical dating methods usually used is Pb-210 dating. The half-life of Pb-210 is about 22.2 years. Therefore, dating is limited to between several and 150 years, and the resolution depends on the accuracy of analysis. Another method is the C-14 method. The half-life of C-14 is about 5600 years. Thus, the dating is from several thousand years to 20,000–30,000 years. There are some other dating methods, such as chronological methods using annual rings in tree trunks, shells, or corals. This dating, especially the use of a tree’s annual rings, is possible to more than several thousand years (3). Consequently, if we want to know about changes in pollution that occurred during and after the industrial revolution, dendrochronological dating is best fitted for this purpose. 5

DIFFERENCE BETWEEN BARK POCKETS AND ANNUAL RINGS AS POLLUTION TIME CAPSULES

There are many reports of heavy metals that accumulate in tree rings corresponding to historical changes in pollution (4). Trees form a growth ring every year. Many publications have analyzed the pollutants in annual rings and calculated the age of the rings, but this presents problems because most atmospheric pollutants that accumulate in tree rings are limited to water-soluble compounds transported by way of soil and roots. Also, there are time lags, sometimes more than 10 years, between soil pollution and the count of annual rings. Most air pollutants in the form of wet and dry depositions on trees accumulate directly on the outer bark. The concentration of pollutants in the annual rings is very low compared with that of the outer bark because of the dispersal that occurs in the process of pollutant transportation through soils and trees. Furthermore, lateral movement of pollutants may occur between adjacent rings (5). Therefore, it is difficult to obtain direct information on air pollution from annual rings, although some historical trends in pollution are reflected in the concentration of pollutants in the annual rings. In the case of bark pockets, there are no time lags, and dispersal of pollutants and translocation of pollutants from bark pockets after encapsulation is considered to be limited because the outermost parts of the two barks in contact with each other consist of the dead tissue in the bark pockets and the bark pocket is not a passageway for water from the roots to the leaves.

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EXAMPLES OF HISTORICAL MONITORING USING BARK POCKETS

For our research using bark pockets for historical monitoring, a 350-year-old Japanese cedar (Cryptomeria japonica) that had suffered typhoon damage and a 226-year-old Japanese cedar were collected, respectively, from Nikko in 1990 and Yakushima Island in 1992. Nikko, located about 100 km north of Tokyo, is famous for its Toshogu shrine (a World Heritage site) which was built about 360 years ago. Also at this time, avenues of C. japonica were planted along the roadside. Yakushima Island, famous for many C. japonica trees older than a thousand years, includes a national park and is also a World Heritage site. The bark pocket collected at Nikko was formed around 1760–1780 (240– 220 years ago based on 2000) and the bark pocket collected at Yakushima was formed around 1786–1809 (214–191 years ago). The concentrations of lead as a pollutant in the bark pockets and outer bark were determined, since the use of leaded gasoline as a fuel for automobiles has increased drastically since World War II. Many countries still use it and lead pollution is still spreading, although manufacturing and sales of leaded gasoline are already banned in Japan. The results of the analysis showed a marked difference between the amount of lead contained in the bark pockets from the Edo era and that in the outer bark, which reflected modern lead pollution. In the case of the cedar trees from the roadside in Nikko, the concentration of lead in modern samples was about 1000 times

FIG. 11 Relative concentration of lead in the bark pockets collected from Nikko and Yakushima.

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higher than that of the Edo era (Fig. 11). In the case of Yakushima Island, a remote island located in southern Japan, the concentration of lead in modern samples was about 10–20 times higher than that of the Edo era (1). In addition to these samples, trees containing bark pockets were collected in the precincts of Muro Temple in Nara prefecture in Japan, about 60 km south of Kyoto and about 37 km south of Nara. The analysis of bark pockets of C. japonica from Muro Temple using laser ablation ICP-MS revealed clearly the historical changes in mercury and lead pollution over a period from 140 to 70 years ago and up to the present time (6). The respective concentrations of lead and mercury in the modern outer bark were about 40 and 4 times higher than those of 140 years ago. 7

NEW PROJECT CONCERNING ATMOSPHERIC CHANGES IN THE WORLD

While environmental pollution and destruction of nature are expanding on a global scale, it is becoming increasingly important to grasp the influence of human activities on a global environment, especially in the world. In the historical monitoring using bark pockets, heavy metals were analyzed as pollutants. However, similar monitoring are possible for other inorganic and organic pollutants. Therefore, we are planning to propose a new project for historical monitoring of atmospheric changes. In addition, we are planning to select the forests for historical monitoring (Time Capsule Forests) throughout the world. The Nakagawa experimental forest, Hokkaido University, located at Latitude 44°44′44″, is the first Time Capsule Forest for this purpose. REFERENCES 1. K Satake, A Tanaka, K Kimura. Accumulation of lead in tree trunk bark pockets as pollution time capsules. Sci Tot Environ 181:25–30, 1996. 2. K Satake, K Kimura, A Tanaka, V Virtanen. Two-hundred-years old shoots of the epiphytic moss Brotherella henonii preserved in a bark pocket of the conifer Cryptomeria japonica. J Bryol 18:815–832, 1995. 3. FH Schweingruber. Tree Rings. Dordrecht: Kluwer Academic, 1988. 4. CF Baes III, HL Ragsdale. Age-specific lead distribution in xylem rings of tree genera in Atlanta, Georgia. Environ Pollut Ser B 2:21–35, 1981. 5. JR Donnelly, JB Shance, PG Schaberg. Lead mobility within the xylem of red spruce seedlings: implications for the development of pollution histories. J Environ Qual 19: 268–271, 1990. 6. K Satake, R Idegawa, M Obata, N Furuta, N. Historical environmental monitoring using bark pockets as pollution time capsules. Proceedings of the Fifth International Conference on the Biogeochemistry of Trace Elements; Vienna 99, Vol. 2. 1074– 1075, 1999.

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13 Tree Rings and Dendroanalysis C. Nabais and H. Freitas Universidade de Coimbra, Coimbra, Portugal

J. Hagemeyer University of Bielefeld, Bielefeld, Germany

1

DENDROANALYSIS: DEFINITION AND FUNDAMENTAL PRINCIPLES

Trees in temperate regions produce annual growth rings (1). Such rings are the result of a rhythmic activity of the cambium. During spring and early summer the cambium is active and new wood is formed, whereas in autumn and winter the cambium is dormant (2). The mineral composition of the wood rings depends partly on mineral uptake of the tree. The extent of mineral uptake may reflect the availability of the elements in the environment surrounding the tree (3,4). This has lead to the idea of using tree rings as retrospective biomonitors of trace metal pollution; the method is known as dendroanalysis (5). To use dendroanalysis as a tool for biomonitoring environmental pollution, some conditions must be meet: (a) a constant relation must exist between the element concentration in the environment and the amount of the element present 367

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in the wood tissue; (b) incorporation of the element into a certain annual ring must occur within a limited period of time; (c) it must be known to which rings the element is transported, either only to the outermost and youngest ring (e.g., in ring-porous trees water conduction takes place primarily in the youngest growth ring) or to a number of older rings; (d) after incorporation in the annual ring no subsequent remobilization or radial transport should occur; (e) the radial distribution pattern of an element should not change its profile over a long period; (f) the radial distribution of a certain element should be similar in different parts of the same tree and in different trees growing in the same area (3,4).

2

LITERATURE REVIEW OF DENDROANALYSIS

A summary of the available publications reporting radial distribution patterns of elements in tree rings is given in Table 1. Some dendroanalytical studies were interpreted as supporting the feasibility of dendroanalysis, while others questioned the method (Table 1). From the published articles cited, 41% state that the method of dendroanalysis to monitor environmental pollution is possible, 26% state that monitoring is not possible, 21% find the method questionable, 8% do not give an opinion on the validation of the method, and 4% conclude that some trace metals can be used to monitor environmental pollution (Table 1). 2.1

Research Supporting Dendroanalysis

In some cases correlations were found between patterns of tree ring heavy-metal contents and temporal records of pollution from other sources, e.g., data of industrial or traffic activities. Several studies have found that Pb concentrations in soils and tree rings declined sharply with increasing distance from the road (6,7). The wood rings of Acer saccharum (8), Quercus nigra (9), and Q. velutina (9) appear to accurately record annual changes in atmospheric metal deposition. Pinus ponderosa growing along rivers with contaminated sediments from the discharge of heavy metal–laden waste showed higher concentrations of Zn and Cd than trees growing at uncontaminated sites (10). The decline of Zn and Cd concentrations in the younger growth rings was explained by the reduction of mining activity and implementation of pollution control technologies that reduced the Zn and Cd pollution in the rivers (10). However, the authors do not exclude the radial translocation of Zn and Cd as a possible influence on the radial distribution patterns. In Quercus robur the radial distributions of Cd and Pb in wood rings growing in the surroundings of a factory showed a good correlation with the calculated emission history (11). However, Cd content in the oak wood continued to increase, even after close-down of the factory. This suggests that Cd was accumu-

Author

Year

Wright & Will (74) Orman & Will (75) Furukawa (76)

1957 1960 1961

Schroeder & Balassa (77) Furukawa (78)

1961 1963

Furukawa (79) Galligan et al. (80) Ault et al. (81) Holtzman (82) Ishizaki et al. (83) McMillin (84) Fossum et al. (85) Janin & Clement (86) Szopa et al. (87) Wardell & Hart (48) Hampp & Ho¨ll (88) Rolfe (89) Ward et al. (90)

1964 1965 1970 1970 1970 1970 1972 1972 1973 1973 1974 1974 1974

Tree genus

Monitoring

Ca, K, Mg, N, Na, P Ca, K, N, P Ca, K, N, P

n.m. n.m. n.m.

Cd, Pb Ca, K, N, P

⫹ n.m.

Ca, K, N, P Mn Pb Pb Cd, Zn Ca, Fe, K, Mg, Mn, Na, P Ca, Mg, Mn Ca, K, Mg, P Pb Ca, Cl, K, Mg, Mn, P, S Pb Pb Pb

n.m. n.m. ? ⫹ ? n.m. n.m. n.m. ⫺ ⫺ ⫺ ? ⫹

369

Pinus Pinus Cedrus, Chamaecyparis, Cryptomeria, Gingko, Larix, Pinus, Taxodium Ulmus Ailanthus, Alnus, Cercidiphyllum, Cinnamomum, Fagus, Fraxinus, Idesia, Lithocarpus, Paulownia, Phellodendron, Prunus, Quercus, Robinia Cryptomeria, Pinus Pseudotsuga Quercus Carya, Quercus, Ulmus Cryptomeria Pinus Pinus Populus Pinus, Quercus Quercus Acer, Robinia, Tilia Acer, Pinus, Quercus Acer, Aesculus, Fraxinus, Platanus, Quercus, Ulmus

Element

Tree Rings and Dendroanalysis

TABLE 1 List of Publications Presenting Original Data of Radial Distribution Patterns of Elements in Tree Stems a

370

TABLE 1 Continued Year

Tree genus

Basham & Cowling (91)

1975

Hall et al. (92) Ricci (93)

1975 1975

Picea, Pinus, Populus, Quercus Fraxinus, Quercus Pinus

Sheppard & Funk (94)

1975

Pinus

Suzuki (95) Barnes et al. (96) Clement & Janin (97) Dollard et al. (98) Faucherre & Dutot (99) Gilboy et al. (5) Lepp (100)

1975 1976 1976 1976 1976 1976 1976

Pillay (101)

1976

Cryptomeria Carpinus, Picea Populus Acer Pinus Ulmus Acer, Fagus, Populus, Tilia, Ulmus Gleditsia, Prunus, Quercus, Syringa, Thuja, Tsuga

Tian & Lepp (102) Tout et al. (103) Ward et al. (104) Herrmann et al. (105) Hincman et al. (106)

1977 1977 1977 1978 1978

Acer, Larix Cedrus, Ulmus Beilschmiedia Quercus Pinus

Kardell & Larsson (107)

1978

Quercus

Element

Monitoring

Mg, P

⫺

Pb Ag, Al, Ba, Br, Ca, Cl, Co, Cr, Fe, Hg, K, Mg, Mn, Na, Rb, V, W, Zn Ag, Au, Co, Cr, Fe, Hg, K, La, Mn, Na, Rb, Sb, Zn Cd, Pb, Zn Pb Ca, K, Mg, Mn, P, Zn Pb Zn K Cu, Pb

? ?

Ag, Al, As, Au, Ca, Cl, Co, Cr, Cu, Fe, Hf, Hg, K, La, Mn, Na, Rb, Zn Pb, Zn Br, Cl, J, K, Mg, Na, Sr Ag Cu, Fe, Mn, Pb, Zn Br, Fe, K, Mn, Na, Pb, W, Zn Cd, Pb

⫹ ⫹ ⫺ n.m. ? ⫹ ? ⫹ ⫹

? ⫺ ? ⫹ ⫹ ?

Nabais et al.

Author

1978 1979 1979 1979

Ulmus Pinus Quercus Pinus

Cl Zn Cd Cd, Cu, Pb, Zn

Valkovic et al. (112)

1979

Not specified

Fergusson et al. (113) Schrimpff (114) Baes & Ragsdale (6)

1980 1980 1981

Robitaille (115) Suzuki (116) Suzuki (117) Isermann (118) Saka & Goring (119) Wickern & Breckle (120) Baes & McLaughlin (121) Baes et al. (122)

1981 1981 1982 1983 1983 1983 1984 1984

Tilia Abies Carya, Liriodendron, Quercus Abies Cryptomeria Pinus Picea, Prunus, Quercus Picea Quercus Pinus Pinus

Ni, Pb, Ti, various element ratios Pb Cd, Cu, Mn, Pb, Zn Pb

Brownridge (123)

1984

Ciepal & Niemtur (124) Greune & Fengel (125)

1984 1984

Acer, Carya, Fagus, Quercus, Robinia Pinus Picea

Kazmierczakowa et al. (126) Legge et al. (127) Ogihara & Katsumo (128)

1984 1984 1984

Fagus Pinus Pinus

⫺ ⫹ ⫹ ⫹ (Cu, Pb) ⫺ (Zn, Pb) ⫹ ⫺ ⫹ ⫹

Cu, Pb, Zn Cd Hg Al, Cd, Cu, Mn, Pb, Zn Ca, Na, S Pb Al, Fe, Mo, Ti, Zn Al, Ca, Cd, Cu, Fe, Mn, Mo, Ti, Zn 137 Cs, 40 K

⫹ ⫹ ⫹ ⫹ n.m. ⫹ ⫹ ⫹

Pb, Zn Al, Ba, Ca, Cl, Cu, Fe, K, Mg, Mn, Na, P, S, Si, Zn Cd, Cr, Fe, Pb, Zn Al, As, Cl, Cu, Mn, P, Si, Zn Al, Cd, Cr, Cu, F, Hg, Ni, Zn

? n.m.

Tree Rings and Dendroanalysis

Tout & Gilboy (108) Dutot & Faucherre (109) Gilboy et al. (110) Symeonides (111)

?

⫺ ? ⫹

371

372

TABLE 1 Continued Year

Tree genus

Taneda et al. (129)

1984

Bauch et al. (130)

1985

Castanea, Cryptomeria, Kalopanax, Prunus, Thujopsis Picea

Bra¨ker et al. (131)

1985

Pinus

Berish & Ragsdale (132)

1985

Carya

Breckle & Schu¨re (133) Kosmus & Grill (134) Matusiewicz & Barnes (135)

1985 1985 1985

Fraxinus, Quercus Aesculus Acer, Picea

Maurer & Peters (136) McLaughlin & Bra¨ker (137) Hagemeyer & Breckle (60) Katayama et al. (138)

1985 1985 1986 1986

Picea Picea Quercus Cryptomeria

Katayama et al. (139) Lim & Cousens (140) Maurer et al. (141) Meisch et al. (142)

1986 1986 1986 1986

Cryptomeria Pinus Picea Fagus

Meisch et al. (143)

1986

Fagus

Element Ca, Cu, Fe, Mg, Mn, Zn

Al, Ca, Cd, Fe, K, Mg, Mn, Pb, Zn Al, Ba, Ca, Cd, Cu, Fe, Mg, Mn, Mo, P, Ti, Zn Al, Cd, Cu, K, Mg, Mn, Ni, Pb, Zn Na Cd, Cu, Pb, Zn Al, As, Ba, Ca, Cu, Fe, Ge, K, Mg, Na, Si, Sr, V, Zn Ca, Pb Fe Cd Al, Br, Ca, Ce, Cl, Co, Cs, Fe, K, La, Mg, Mn, Na, Rb, Sc, Sm, Ti, V Cs, 137 Cs, K, 40 K, Na, Rb Ca, K, Mg, N, P Cl, N, F, P, S Al, Ca, Co, Cr, Cu, K, Mg, Mn, Na, Ni, Pb, Zn Al, Ca, Fe, Mg, Mn, Zn

Monitoring n.m.

? ? ⫹ ⫺ ? ? ⫹ ⫹ ⫺ ⫺ ⫺ ⫺ ? ⫹ ⫹

Nabais et al.

Author

1986

Baes & McLaughlin (66) Guyette & McGinnes (145) Hofmann & Born (146) Maclauchlan et al. (147) McLenahen et al. (148) Nagj et al. (149)

1987 1987 1987 1987 1987 1987

Castanea, Cryptomeria, Kalopunax, Prunus, Thujopsis Pinus Juniperus Fraxinus Alnus Liriodendron Pinus

Okada et al. (150)

1987

Cryptomeria

Queirolo & Valenta (151) Tru¨by & Zo¨ttl (152) Arp & Manasc (153)

1987 1987 1988

Quercus Betula, Pseudotsuga Picea

Chigira et al. (154)

1988

Cryptomeria

Kohno et al. (155)

1988

Kyncl et al. (156) Lukaszewski et al. (157) Okada et al. (158) Ragsdale & Berish (159) Scherbatskoy & Matusiewicz (160)

1988 1988 1988 1988 1988

Chamaecyparis, Cryptomeria Picea Pinus Cryptomeria Carya Acer, Picea

Tendel & Wolf (161)

1988

Pinus

Ca, Cu, Fe, K, Mg, Mn, Na, Zn

⫺

Fe, Zn Al, Cu, Fe, Mn, Zn Pb Ca, Cu, Fe, Pb, Sr, Zn Ca, P Ca, Cr, Cu, Fe, K, Mn, Pb, Rb, S, Ti, V, Zn Al, Br, Ca, Cl, Cr, K, Mg, Mn, Na, Rb Cd, Pb Cd, Pb, Zn Al, Ca, Cr, Cu, Fe, K, Mg, Mn, Ni, P, Zn Ca, Cs, 137 Cs, K, Na, Rb, 90 Sr

⫹ ⫹ ⫹ ⫹ ⫹ ?

137

Cs

⫹ ⫺ ? ⫹(Sr) ⫺(Cs) ⫺ ⫺ ⫺ ⫺ ⫹ ⫹ ⫹

373

Ca, Cu, Fe, Mg, Mn, Pb Cd, Cu, Zn Cl, K, Mg, Mn Pb Al, As, Ba, Ca, Cu, Fe, Ge, K, Mg, Mn, Na, Si, Sr, V, Zn Al, Ca, Cd, Cu, Fe, K, Mg, Mn, Ni, P, Pb, S, Zn

⫺

Tree Rings and Dendroanalysis

Taneda et al. (144)

TABLE 1 Continued

Tru¨by (162)

1988

Tru¨by & Zo¨ttl (163) Bondietti et al. (164) Frelich et al. (165)

1988 1989 1989

Abies, Pseudotsuga, Quercus Abies, Pseudotsuga Picea, Tsuga Acer, Pinus

Guyette et al. (166) Hagemeyer et al. (167)

1989 1989

Juniperus Fagus

Ha¨sa¨nen & Huttunen (168)

1989

Pinus

Helmisaari & Siltala (169)

1989

Pinus

Ilgen & Nebe (170) Long & Davis (171)

1989 1989

Picea Quercus

McClenahen et al. (55) Queirolo (172) Queirolo et al. (173) Aberg et al. (174) Bondietti et al. (13) Hall et al. (175) Jordan et al. (176) Kairiukstis & Kocharov (177) Momoshima & Bondietti (30) Okada et al. (40)

1989 1989 1989 1990 1990 1990 1990 1990 1990 1990

Liriodendron Quercus Quercus Picea, Pinus Abies, Betula, Picea, Tsuga Pseudotsuga Pinus Pinus Picea Cryptomeria

Tree genus

Element

Monitoring

Pb

⫺

Cd, Pb, Zn Al, Ca, Mg, 90 Sr Ca, Fe, K, Mg, Mn, P, Pb, S, Zn Mo, S Ca, Cd, K, Mg, Na, Ni, Pb, Zn Al, B, Ca, K, Mg, Mn, Mo, N, P, Rb, S, Zn Al, B, Ca, K, Mg, Mn, N, P, Zn As, Cd, Cu, N, Pb, S, Zn Al, B, Ba, Ca, Cd, Co, Cu, Cr, Fe, K, Mg, Mn, Na, Ni, P, Pb, Si, Sr Ca, K, Mn, P, Sr Cd, Cu, Pb, Zn Cd, Pb Ca Ca, Mg, Mn, Sr Ce, La, Nd, Sm Al, Cd, Cu, Mn, Pb, Zn K Ca, P, cation ratios Ca, Cl, K, Mg, Mn, Na, Rb

⫺ ⫹ ? ⫹ ⫺ ? n.m. ⫹ ⫺

? ? ? ⫹ ? ⫹ ? ⫹ ⫹ ⫺

Nabais et al.

Year

374

Author

Quercus

Pernestal et al. (178)

1990

Picea

Queirolo et al. (179) Tru¨by & Lindner (180) Tru¨by & Zo¨ttl (181) Vroblesky & Yanosky (182) Zech et al. (183) Gilfrich et al. (184)

1990 1990 1990 1990 1990 1991

Quercus Picea Picea Liriodendron Fagus Quercus, Sassafras

Guyette et al. (185) Kennedy & Bergeron (186)

1991 1991

Ohmann & Grigal (187)

1991

Juniperus Acer, Fraxinus, Populus, Thuja, Tsuga Abies, Acer, Pinus, Populus

Pernestal et al. (188)

1991

Picea

Ritters et al. (189)

1991

Stewart et al. (7) Zayed et al. (190) Chun & Hui-Yi (41)

1991 1991 1992

Abies, Acer, Picea, Pinus, Populus Dacrycarpus Picea Pinus, Quercus

Ferretti et al. (191) Hagemeyer et al. (192) Hantemirov (73) Vroblesky et al. (193) Wickman (194)

1992 1992 1992 1992 1992

Pinus Fagus Pinus Liriodendron Picea, Pinus

Al, Ba, Br, Ca, Cl, Cs, Cu, K, Mg, Mn, Na, Rb Ca, Cl, Cu, Fe, K, Mn, S, Sr, Zn Cd, Cu, Pb, Zn Mn Cd, Pb, Zn Cl, Fe Mg, S Ca, Cl, Co, Cu, Fe, K, Mn, Ni, Ti, Zn Cd, Pb Mn Al, B, Ca, Cd, Cr, Cu, Fe, K, Mg, Mn, N, Na, Ni, P, Pb, S, Zn Ca, Cr, Cu, Fe, K, Mg, Mn, Zn Al, B, Ca, Cu, Fe, K, Mg, Mn, N, P, S, Zn Cd, Cu, Mn, Pb, Zn Al Al, As, Ca, Cd, Cr, Cu, Fe, K, Mg, Mn, Ni, P, Pb, Ti, V, Zn Cl, P, S Cd, Pb, Zn Ca, Cu, K, Mg K Ca, Mg, Mn, Sr, 87 Sr/ 86 Sr

⫺ ? ? n.m. ⫺ ⫹ ⫹ ? ⫹ ⫹ ?

⫹ ? ⫹ ? ⫹/⫺ ⫹ ⫺ ⫹ ⫺ ?

375

1990

Tree Rings and Dendroanalysis

Okada et al. (44)

376

TABLE 1 Continued Year

Tree genus

Yanosky & Vroblesky (195)

1992

Zayed et al. (196) Ferretti et al. (197)

1992 1993

Liquidambar, Liriodendron, Prunus, Quercus, Sassafras Picea Pinus

Hemmann (198)

1993

Pinus

Katayama et al. (199) Kudo et al. (200)

1993 1993

Cryptomeria Cryptomeria

Lukaszewski et al. (201) Okada et al. (45)

1993 1993

Okada et al. (46)

1993

Wenk (202) Balk & Hagemeyer (69) Dittmar & Zech (203) Forget et al. (204) Hagemeyer et al. (15) Hoffmann et al. (205) Hofmann (206)

1993 1994 1994 1994 1994 1994 1994

Populus Populus, Abies, Cedrus, Chamaecyparis, Larix, Metasequoia, Pinus Castanea, Hovenia, Lindera, Phellodendron, Quercus, Sorbus, Stewartia, Zelkova Picea Quercus Fagus Picea Fagus Pinus, Populus Fraxinus

Element

Monitoring

Ni

⫹

Al, Ca, Cu, Fe, Mg, Mn, Zn Al, Ca, Cd, Cu, Fe, K, Mg, Mn, Na, Pb, Rb, Zn Al, B, Ca, Cd, Cu, Fe, K, Mg, Mn, Na, Zn As, Cu, K 137 Cs, 40 K, 239 Pu239, 240 Pu

⫺ ⫹

Cd, Cu, Pb, Zn Mn, Na, Rb, Sm, Zn

⫺ ⫹ ⫹(Pu) ⫺(Cs, K) ⫺ ⫺

Al, Ba, Br, Ca, Ce, Cl, Cs, K, La, Mg, Mn, Na, Rb, Sm, Sr, Zn

⫺

N, S Cd, Pb Mg, Mn, S Mn Ni Mg, Al/Mg, Pb/Mg Pb

⫹ ⫺ ⫹ ⫺ ⫺ ⫹ ⫹

Nabais et al.

Author

1994

Momoshima et al. (208) Myre & Camire (209) Tru¨by (210)

1994 1994 1994

Aoki et al. (211) Cote & Camire (212) Eklund (11) Garrec et al. (213) Hagemeyer (61) Hagemeyer & Scha¨fer (19) Hagemeyer & Shin (68) Kardell (214) Momoshima et al. (12)

1995 1995 1995 1995 1995 1995 1995 1995 1995

Acer, Carya, Fagus, Liriodendron, Tsuga, Ulmus, Picea, Pinus Abies, Cryptomeria Larix Abies, Fagus, Picea, Pinus, Pseudotsuga, Quercus Cryptomeria Acer Quercus Abies, Cryptomeria, Picea Quercus Fagus Pinus Pinus, Quercus Cryptomeria

Shortle et al. (215) Tru¨by (216) Li et al. (217) Poulson et al. (218) Suzuki (219) Yanosky et al. (220) Eklund et al. (221) Latimer et al. (222) Watmough & Hutchinson (223)

1995 1995 1995 1995 1995 1995 1996 1996 1996

Abies, Picea Picea, Pinus, Quercus Picea Tsuga Cryptomeria Taxodium Quercus Taxodium Acer

137

Cs,

90

Sr

Ca, K, Mg, P Ca, K, Mg, Mn, P, Zn Ca, Cd, Cu, K, Mg, Mn, P, Pb, Zn 137 Cs, 90 Sr Al, Ca, K, Mg Cd, Pb Cs, K, Pu Cd Cd, Pb, Zn Cd, Pb Cd, Cu, Pb, Zn Ba, Ca, Cs, K, Mg, Mn, Na, P, Rb, Sr, Ca Cd, Cu, Pb, Zn Hg N Cd, Cu, Hg, V Cl Cd, Pb Pb, Zn Ca, Cd, Cu, K, Fe, Mg, Mn, Ni, Pb, Sr, Zn

⫹(Sr) ⫺(Cs) ⫺ ⫹ ⫺ ⫺ ? ⫹ ? ⫺ ⫺ ⫺ ⫺ ⫹

Tree Rings and Dendroanalysis

Momoshima & Bondietti (207)

⫹ ⫺ ? ⫹ ⫹ ⫹ ⫹ ⫹ ⫹

377

378

TABLE 1 Continued Year

Garbe-Schønberg et al. (14) Jonsson et al. (224) Oliveira et al. (225)

1997 1997 1997

Betula, Pinus Quercus Pinus

Watmough et al. (226) Hofmann et al. (227) Horn et al. (20) Larsson & Helmisaari (228)

1997 1998 1998 1998

Acer Picea Pinus

Prohaska et al. (22)

1998

Picea

Schaumloffel et al. (10) Watmough et al. (8) Dewalle et al. (229)

1998 1998 1999

Penninckx et al. (230) Watmough (231) Watmough & Hutchinson (67)

1999 1999 1999

Pinus Acer Acer, Betula, Fagus, Larix, Liriodendron, Picea, Prunus Fagus Abies, Acer, Picea Abies

Anderson et al. (9)

2000

Quercus

Tree genus

Element Al, Cu, Pb Cd, Pb Al, B, Br, Ca, Co, Cs, Fe, K, Mg, Mn, Na, P, Rb, Zn Cd, Mg, Mn, Pb Al, Ba, Ca, La, Mn, P Pb Ca, Cu, Fe, K, Mn, Ni, Rb, Sr, Zn Al, Ba, Ca, Cd, Co, Cr, Fe, Mg, Mn, Pb, Sr, Zn Cd, K, Zn Ca, Cd, Cr, Cu, Mn, Pb, Zn Al, Ca, Mg, Mn

Ca, K, Mg, Mn, N, P Ca, Cd, Cu, Pb Al, Ca, Cd, Cu, Mg, Mn, Ni, Pb, Zn Pb

Monitoring ⫺ ⫹ ? ⫹ ⫹ ⫺ ⫹/⫺ ⫺ ⫹ ⫹ ⫹/⫺

? ⫹/⫺ ⫹ ⫹

a The column ‘‘Monitoring’’ summarizes the authors’ conclusions about the value of dendrochemistry in biomonitoring: (⫹) monitoring is possible; (⫺) monitoring is not possible; (?) method questionable; (n.m.) not mentioned or monitoring was not intended. Source: Updated from Refs. 3 and 4.

Nabais et al.

Author

Tree Rings and Dendroanalysis

379

lated in the soil profile and subsequently mobilized, absorbed by the roots, and incorporated in the wood (Fig. 1) (11). Japanese cedar or sugi (Cryptomeria japonica D. Don) shows a decline in Japan probably due to acidic deposition (12). Acidic deposition in the soil can change the availability of elements for uptake by plants. To trace back the acidic deposition history in a forest it is necessary to understand the distribution and mobility of cations in the xylem. Distributions of cations that are immobile in annual rings can be used as a temporal record of nutrient availability (13). In the annual rings of sugi the radial distribution of 90 Sr was related to the cumulative deposition of fallout from nuclear weapon tests, while the radial distribution of 137 Cs was unrelated (12). This suggests that Sr 2⫹ is immobile while Cs ⫹ is mobile in the horizontal direction in wood. The radial distribution of Sr 2⫹ in sugi could be used as a chronological index of nutrient availability (12).

FIG. 1 Radial distribution of Cd in stem wood of Quercus robur adjacent to a factory and from an unpolluted reference site. Estimated emissions of Cd from the factory are also presented (From Ref. 11.)

380

2.2

Nabais et al.

Research Questioning Dendroanalysis

The feasibility of dendroanalysis is not beyond doubt and in several studies the basic assumptions of the method have been questioned. The Cu, Al, and Pb levels in rings of Pinus silvestris and Betula sp. growing near nonpolluted areas were similar to those from trees in the immediate surroundings of industrial pollution sources, although metal levels in the soil may be up to 1000 times higher in the contaminated site (14). The temporal profiles in tree rings could not be linked to the pollution history of the area. Seasonal variations of Ni concentrations in stem wood of Fagus sylvatica were detected (15). Highest levels of Ni were determined during the spring mobilization period in April (15). Trace element cations in the xylem tissue can be mobile in the xylem sap or bound by electrostatic force to fixed negative charges in the xylem vessel walls (16). Binding to fixed negative charges is a reversible process. During the annual cycle of growth in deciduous plants, the ionic composition or the pH of the xylem sap can change and ions can be remobilized and transported to other parts of the tree (17,18). These processes could explain the variations of Ni concentrations in the wood of Fagus sylvatica (15). Therefore,

FIG. 2 Radial distribution of 206 Pb/ 207 Pb in growth rings and bark of Picea abies in Germany. The isotopic ratio of industrial lead used in Germany since 1980 are below 1.16, whereas before it was about 1.17. Geogenic lead in Germany has 206 Pb/ 207 Pb above 1.18. (From Ref. 20.)

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seasonal variations of Ni concentrations in wood preclude the use of radial Ni distributions for the reconstruction of pollution levels in the past (19). The industrial and traffic lead emissions show very different isotopic signatures (206 Pb/ 207 Pb) from those of local soils. Hence, the isotopic rations allow differentiation between these sources (20). It would be expected to find different isotopic signatures of Pb in wood rings from the year 1950 onward with an increase in environmental pollution. The ratios of 206 Pb to 207 Pb in growth rings and bark from Norway spruce (Picea abies) showed a large change from geogenic, or pedogenic, lead isotope values (1930–1937) to more anthropogenic industrial ratios only at the end of the trees’ lifetime in 1990 (Fig. 2) (20). However, from 1943 to 1976 the isotopic ratio remains constant, although during this period an increase in industrial and traffic emissions was observed, with a change in 206 Pb/ 207 Pb (20). Therefore there was no conservation of past trace metal pollution in growth rings of Picea abies. 3

TRANSPORT AND DEPOSITION OF TRACE METALS IN THE XYLEM

Trees are not passive recorders of the external environment, and the application of dendroanalysis to environmental monitoring requires an understanding of the factors that influence the mobility of trace metals in the xylem (21). The accumulation of elements in annually formed tree rings is controlled by such factors as the cation binding capacity of wood, transformation of sapwood into heartwood, and processes of radial translocation of elements in the tree stem (22). 3.1

Cation Binding Capacity of Wood

Trace metals that enter roots are transported in the xylem with the transpiration stream. The xylem sap in a tree is the liquid that flows in the apoplast, mainly in the conducting elements of the wood (23). Wood is mainly composed of cellulose, pectins, and lignin (24). The cellulose microfibrils are embedded in a gel of pectic polysaccharides, the major component of which is polygalacturonic acid. These acids bear negatively charged carboxyl groups, which are mainly responsible for the overall negative charge of the cell wall (25). The amount of nondiffusible negative charges in the apoplast is normally quantified by the cation exchange capacity (CEC) (26). Most elements do not follow simple diffusion or mass flow kinetics from the soil to the stem (27). Interactions of cations with nondiffusible anions of the cell wall of the xylem vessels lead to a separation of cation transport from water flow (28). The xylem operates as an ion exchange column that has the potential to impede movement up the bole for months or even years (29,30). Changes in

382

Nabais et al.

wood chemistry sampled at breast height may therefore correspond to changes in soil chemistry that occurred several years earlier. The amount and proportion of ions retained in the xylem depends on the ionic charge density of the ions (Ca 2⫹ ⬎ K ⫹), the activity of competing cations, the charge density of the nondiffusible anions, and the pH of the xylem sap, which may vary from 5 to 7 (26,31). Binding to mobile organic molecules (e.g., carboxylic acids, amino acids, peptides) will also alter the charged state of cations and reduce their affinity for fixed negative charges in cell walls (32). Consequently, the transport rate of di- or trivalent cations is enhanced significantly by complexation of the cation (33). Most of the cations retained in the cell walls are easily exchangeable (28). However, cell walls may also contain multivalent cations precipitated or irreversibly bound to polyuronic acids and other ligands (34). Cations that are relatively mobile in the xylem would obscure the time course changes in the tree’s environment (13). The radial distributions of elements in stem wood depend partly on the cation exchange capacity of the tissue. In stem wood of Picea rubens, the density of fixed negative charges in cell walls is higher in the stem center (30). Most trees show a decline in cation concentration with increasing tree age, which may be linked more to the binding exchange properties in the woody tissue than to the uptake of trace elements from the soil (22). The variation of the CEC with the distance from the stem center suggests that functional groups responsible for cation binding change with time (12). As the stem radius increases, the tracheids are longer with a consequent wider distribution of the negative charges (30). Element concentrations within an individual annual ring may strongly depend on the stem side. In Picea abies element concentrations within a single tree ring varied with the direction by more than 60% (22). This variation was explained by the changed proportion of compression wood at the various sides of the tree stem (22). Compression wood, formed only in gymnosperms, develops in response to unbalanced mechanical pressure (35). The growth rings are wider with higher density and appear darker due to the enhanced lignin and reduced cellulose content. As 60–70% of the cations bind to hydroxyl groups, the ligninrich compression wood provides more binding sites than the normal wood (22). An increased concentration of elements is expected in compression wood due to these differences in wood chemistry. 3.2

Effects of the Sapwood/Heartwood Boundary on Trace Element Distributions

In many tree species heartwood formation is obvious from the color change of the wood. This aging process, leading to cell death and the decomposition of

Tree Rings and Dendroanalysis

383

storage material, is also associated with an increase in content of extractives (36). Some of these heartwood extractives, like phenolics and terpenoids, function as decay inhibitors (37). Heartwood polyphenols are synthesized in situ at the sapwood-heartwood boundary from the breakdown of starch or from soluble sugars (36). During the process of heartwood formation the moisture content of wood decreases, the ray parenchyma cells die, stored starch in these cells disappears, lignification begins much earlier, and the wood is impregnated with extractives (38,39). Many authors observed that the radial distribution patterns of trace elements in tree stems are influenced by the location of the border between sapwood and heartwood (40,41). Toxic elements like Cd and Pb showed low concentrations near the cambium and increased levels toward the stem center. A translocation of toxic organic substances through the rays into the heartwood was suggested as a possible mechanism of sapwood detoxification (42). Nutrient concentrations generally decrease with increasing distance from the pith in heartwood (43). The pattern is disrupted at the heartwood-sapwood boundary, where the nutrient concentration in the sapwood is much higher than the concentration in the youngest heartwood (40,43–46). This discontinuity across the heartwood-sapwood boundary indicates the retranslocation of nutrients from newly forming heartwood to fully functional sapwood (43,47). Wardell and Hart (48) showed that P, K, Mn, Mg, and Ca were retranslocated out of the senescing sapwood via ray cells. Because of the high volume of stem wood in a forest, nutrient retranslocation from forming heartwood to functional sapwood can have a significant effect on intrasystem nutrient cycling. Andrews and Siccama (49) reported nutrient retranslocation in Chamaecyparis thyoides and estimated that 80% of the Ca and 78% of the Mg required for new wood growth was recovered from dying sapwood. The degree of retranslocation of nutrients from senescing sapwood has been shown to be inversely related to soil nutrient availability (43). Once metals enter trees, they are not necessarily restricted to tree rings formed during the current year. In coniferous wood (e.g., Picea) and diffuse porous wood of angiosperms (e.g., Fagus, Acer, Populus), the water movement passes through several annual rings in the outer part of the stem—the sapwood. In an 18-year-old spruce tree (Picea abies) 10–12 tree rings were involved in water transportation (50). Therefore, the element accumulation of an individual ring is not the result of the element uptake in the year when this ring was formed. In ring porous wood (e.g., Quercus, Ulmus, Fraxinus), water movement is mainly confined to the outermost and youngest annual ring (51). In such trees the transport of minerals from the roots reaches mainly the current year’s ring. These conditions are more favorable for an element chronology in stem wood (3).

384

3.3

Nabais et al.

Radial Translocation

The ray parenchyma of wood is specialized for the temporary storage and remobilization of organic and inorganic reserve material (52). The rays have ‘‘contact cells’’ with large pits to the vessel elements and central rows of ‘‘isolation cells’’ without conspicuous pits to the vessel elements. The isolation cells play an essential role in radial transport (53). The minimum radial flux of sugars through the tangential walls in the isolation cells of Populus ⫻ canadensis was 80.7 pmol cm ⫺2 s ⫺1 (52). Donnelly et al. (54) reported that lateral movement of Pb in Picea rubens seedlings occurred both inward and outward. Trace metals may differ in their mobility between tree rings. In tulip trees (Lirodendron tulipifera), lateral movement is significant for such elements as Mn, K, P, Ca, Sr, and Zn but less so for Al, Si, Fe, and Cu (55). However, to our knowledge, no investigations were performed on the storage capacity, mobilization, and translocation rates of trace metals in ray parenchyma cells. In addition to the radial transport of carbohydrates along rays from the phloem into the xylem and vice versa, exchange seems to occur between the rays and the vessels through the contact cells (56). The contact cells differ in their

FIG. 3 Radial distributions of Cd in stem wood of Quercus robur sampled in 1983 and in 1994. The arrows indicate the transition between sapwood and heartwood at the time of sampling. (From Ref. 61.)

Tree Rings and Dendroanalysis

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anatomy from the other ray cells. They are connected with vessels by enlarged pits, which are probably involved in the exchange of solutes between vessels and rays (56). The pits of the contact cells in rays of Betula populifolia and Populus tremuloides showed a high phosphatase activity (57). The enzymatic activity in the contact cells began in spring, when the starch started to break down into sugars in the ray cells and ended when the leaves were fully developed. During this period, sugars could be found in the xylem sap, suggesting that the contact cells of the rays were involved in an active secretion of sugars into vessels (57). The ray cells can play an important role in the process of heartwood formation, through the production of tannins (58). Some authors suggested that the ray cells at the heartwood-sapwood boundary have an increased metabolic activity (59). Peaks of trace metals were observed at the heartwood-sapwood boundary (40,44–46,60). These peaks were suggested to result from centripetal movements of substances in the sapwood (15). This centripetal transport might serve to detoxify the live part of the stem and dispose of toxic substances in the heartwood (61). Probably the ray cells have an important role in the transport of trace metals to the heartwood. The heartwood-sapwood boundary moves with time and the peaks of trace metals follow the location of that boundary (Fig. 3) (61). Therefore, the current locations of these peaks are not related to pollution events in the tree’s environment. 4

‘‘BIOMONITO-RING?’’

Although the idea of dendroanalysis is fascinating, many problems still remain to be solved before the potential of the method can be realistically assessed (62). Some important factors merit further investigation, e.g., the possible pathways of metal uptake (via roots, foliar uptake, absorption through the stem surface) and their relative importance in the accumulation of trace metals in tree rings; the chemical binding forms of trace metals in wood; and the mobility and transport of trace metals in wood rings (63,64). Some authors state that in trees growing in more acidic soils, with high availability of trace metals, uptake by roots is the major pathway by which metals enter into tree rings (21,65). However, in less acidic soils direct uptake through bark and leaves can also be an important pathway to absorb trace metals (21,65). These metals are then translocated to the outermost xylem ring (66). According to Watmough and Hutchinson (67), metals entering via this pathway are more likely to accurately record changes in atmospheric deposition with minimal radial translocation. However, the actual mechanism and extent of metal uptake through the bark and leaves into tree rings have not been determined. Even if trace metals are incorporated in the recent tree rings through bark or leaves, radial translocation is possible.

386

Nabais et al.

One way to assess the mobility of elements in wood is to determine the strength of their chemical binding forms by using several extractants with different binding strength to metal cations (68). The efficiencies of water, CaCl 2, acetic acid, malic acid, citric acid, and Na 2 EDTA in extracting Cd and Pb from wood of Pinus sylvestris and Quercus petraea were investigated (68,69). Three fractions of Cd and Pb in wood were distinguished: water-soluble, exchangeable, and nonexchangeable. The mobile fraction of the metals is dissolved in the xylem sap and is extractable with water. In wood of beech trees (Fagus sylvatica), less than 1% of the total Cd content was dissolved in the xylem sap (18). The less mobile fraction of the metal cations is bound to fixed negative charges in the xylem vessel walls (28). The fraction of Cd and Pb extracted with solutions of organic acids, Na 2 EDTA, and CaCl 2 includes cations bound to fixed negative charges in the xylem vessel walls. Extracted amounts of Cd increased in the order acetic acid ⬍ malic acid ⬍ citric acid ⬍ EDTA. The number of carboxylic groups of such ligands increases from 1 (acetic acid) to 4 (EDTA) (68). The stability constants of complexes of these ligands formed with Cd and Pb increase in the same order (70). Such ligands compete for cations with fixed negative charges in the wood. They extract different fractions of the bound Cd and Pb. In the case of extraction with CaCl 2 solution, only ion exchange without metal complexation is involved. The Ca-exchangeable fraction contained substantial amounts of Cd and Pb. The nonextractable fraction comprised 20% of total Cd and 50% of total Pb (68). Elements of this fraction can be precipitated as insoluble phosphates, or immobilized in crystal structures, like oxalates. A high percentage of Cd and Pb in wood tissue is mobile or can be remobilized. The seasonal changes of the organic acid composition of the xylem sap may affect mineral concentrations in wood by varying cation transport and remobilization (71,72). Therefore, radial distribution patterns of trace elements in wood can probably change their profile with time. The total content of the chemical elements in rings cannot be used for retrospective biomonitoring (73). Only the chemical forms of an element that are strongly connected with the xylem can eventually be used in pollution reconstruction. A few conditions under which dendroanalytical biomonitoring may have a better chance to produce useful results can be noted: 1. When using an element of low mobility in plants, e.g., Pb. 2. When the selected tree species have a ring-porous wood, in which most of the xylem sap goes through one or two rings; 3. The time passing between emission of an element into the environment

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and its uptake by the tree can differ and this possible time shift should be taken into account when interpreting the results. 4. Only long-term trends in the history of environmental pollution should be considered in radial distribution patterns. 5. Places with high deposition of pollutants and abrupt changes in pollution are the most suitable for dendroanalysis. 5

CONCLUSIONS

A fundamental assumption of the dendroanalytical method is the stability of the mineral distribution patterns, i.e., once the elements are stored in wood rings no significant mobility should occur. Although some studies presented good correlations between radial distributions of heavy metals in tree rings and temporal records of pollution from industry or traffic, others failed in using dendroanalysis as a chronological record of pollution. Radial element distributions can change with time. Mobile portions of elements in wood indicate the possibility of remobilization and retranslocation. Therefore, tree ring analysis does not provide a reliable record of the pollution history. Nevertheless, the method can provide information about transport and storage of elements in trees. REFERENCES 1. HC Fritts. Tree Rings and Climate. London: Academic Press, 1973. 2. FH Schweingruber. Trees and Wood in Dendrochronology. Berlin: Springer-Verlag, Berlin, 1993. 3. J Hagemeyer. Monitoring trace metal pollution with tree rings: a critical reassessment. In: B Markert, ed. Plants as Biomonitors. Weinheim: VCH, 1993, pp. 541– 563. 4. J Hagemeyer. Variabillita¨t der Elementverteilungen im Stammholz von Ba¨umen. Dissertationes Botanicae 279:1–84, 1997. 5. WB Gilboy, RE Tout, NM Spirou. Dendroanalysis: the study of trace elements in tree rings. Proceedings ERDA X- and Gamma-Ray Symposium, Ann Arbor, 1976, pp. 165–165. 6. CF Baes, HL Ragsdale. Age-specific lead distribution in xylem rings of three tree genera in Atlanta, Georgia. Environ Pollut 2:21–35, 1981. 7. C Stewart, DA Norton, JE Fergusson. Historical monitoring of heavy metals in kahikatea ring wood in Christchurch, New Zealand. Sci Tot Environ 105:171–190, 1991. 8. SA Watmough, TC Hutchinson, EPS Sager. Changes in tree ring chemistry in sugar maple (Acer saccharum) along an urban-rural gradient in southern Ontario. Environ Pollut 101:381–390, 1998.

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151. F Queirolo, P Valenta. Trace determination of Cd, Cu, Pb and Zn in annual growth rings by differential pulse anodic stripping voltammetry. Fresenius Z Anal Chem 328:93–98, 1987. 152. P Tru¨by, HW Zo¨ttl. Atmogene und geogene Komponenten in Schwermetallhaushalt von Waldba¨umen. KfK PEF-Berichte 12:333–348, 1987. 153. PA Arp, J Manasc. Red spruce stands downwind from a coal-burning power generator: tree-ring analysis. Can J For Res 18:251–264, 1988. 154. M Chigira, Y Saito, K Kimura. Distribution of 90 Sr and 137 Cs in annual tree rings of Japanese Cedar, Cryptomeria japonica D. Don. J Radiation Res 29:152–160, 1988. 155. M Kohno, Y Koizumi, K Okumura, I Mito. Distribution of environmental cesium137 in tree rings. J Environ Radioact 8:15–19, 1988. 156. J Kyncl, J Dobry, J Jambor. Metal content in the spruce wood (Picea abies (L.) Karst.) in the emission area ‘‘Jizerske Hory’’. Drevarsky Vyskum 119:1–24, 1988. 157. Z Lukaszewski, R Siwecki, J Opydo, W Zembrzuski. The effect of industrial pollution on zinc, cadmium and copper concentration in the xylem rings of Scot’s pine (Pinus sylvestris L.) and in the soil. Trees 2:1–6, 1988. 158. N Okada, M Sato, Y Katayama, T Nobuchi, Y Ishimaru, H Yamashita, A Aoki. Trace elements in the stems of trees II. Influence of age and vertical position on radial distribution in sugi (Cryptomeria japonica). Mokuzai Gakkaishi 34:874–880, 1988. 159. HL Ragsdale, CW Berish. The decline of lead in tree rings of Carya spp. in urban Atlanta, GA, USA. Biogeochemistry 6:21–29, 1988. 160. T Scherbatskoy, H Matusiewicz. Chemistry of annual rings of red spruce and sugar maple in Vermont. Agricultural Experiment Station University of Vermont Research Report 53, 1988, pp. 1–9. 161. J Tendel, K Wolf. Distribution of nutrients and trace elements in annual rings of pine trees (Pinus sylvestris) as an indicator of environmental changes. Experientia 44:975–980, 1988. 162. P Tru¨by. Bleiverteilungen in Waldba¨umen unterschiedlich belasteter Standorte. Angew Bot 62:93–104, 1988. 163. P Tru¨by, HW Zo¨ttl. Schwermetallverteilungen in Waldba¨umen geogen unterschiedlich belasteter Standorte. Kernforschungsanlage Karlsruhe PEF-Berichte 35:231– 243, 1988. 164. EA Bondietti, CF Baes, SB McLaughlin. Radial trends in cation ratios in tree rings as indicators of the impact of atmospheric deposition on forests. Can J For Res 19: 586–594, 1989. 165. LE Frelich, JG Bockheim, JE Leide. Historical trends in tree-ring growth and chemistry across an air-quality gradient in Wisconsin. Can J For Res 19:113–121, 1989. 166. RP Guyette, BE Cutter, GS Henderson. Long-term relationships between molybdenum and sulfur concentrations in red cedar tree rings. J Environ Qual 18:385–389, 1989. 167. J Hagemeyer, B Kamradt, H Scha¨fer, K Schlagintweit, L Verlage, SW Breckle. Saisonale Schwankungen der Elementgehalte und Histologie des Kambiums von Buchenholz in Nordrhein-Westfalen. Allg Forstz 29–30:769–771, 1989.

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168. E Ha¨sa¨nen, S Huttunen. Acid deposition and the element composition of pine tree rings. Chemosphere 18:1913–1920, 1989. 169. HS Helmisaari, T Siltala. Variation in nutrient concentrations of Pinus sylvestris stems. Scandinavian J For Res 4:443–451, 1989. 170. G Ilgen, W Nebe. Jahrringchronologische Differenzierung chemischer Elemente im Holza¨lterer Fichten Biol Rundschau 27:237–247, 1989. 171. RP Long, DD Davis. Major and trace element concentrations in surface organic layers, mineral soil, and white oak xylem downwind from a coal-fired power plant. Can J For Res 19:1603–1615, 1989. 172. F Queirolo. Leistungsfa¨higes voltammetrisches Analysenverfahren zur Spurenbestimmung der Schwermetalle Cd, Pb, Cu und Zn in kleinen Holzmengen. Berichte der Kernforschungsanlage Ju¨lich Nr. 2333 (Dissert. Univ. Bonn), 1989. 173. F Queirolo, P Valenta, S Stengen, SW Breckle. Accumulation of heavy metals in oak wood from polluted regions. In: HW Georgii, ed. Mechanisms and Effects of Pollutant-Transfer into Forests. Dordrecht: Kluwer Academic, 1989, pp. 193–202. 174. G Aberg, G Jacks, T Wickman, PJ Hamilton. Strontium isotopes in trees as an indicator for calcium availability. Catena 17:1–11, 1990. 175. GS Hall, DK Yamaguchi, TM Rettberg. Multielemental analyses of tree rings by inductively coupled plasma mass spectrometry. J Radioanal Nucl Chem Lett 146: 255–265, 1990. 176. DN Jordan, LM Wright, BG Lockaby. Relationship between xylem trace metals and radial growth of loblolly pine in rural Alabama. J Environ Qual 19:504–508, 1990. 177. L Kairiukstis, GE Kocharov. Measuring the chemical ingredients in tree rings. In: ER Cook, LA Kairiukstis, eds. Methods of Dendrochronology. Dordrecht: Kluwer Academic, 1990, pp. 229–232. 178. K Pernestal, HK Li, B Jonsson. Analysis of trace element content in tree cores by PIXE. Nucl Instr Meth Phys Res 49:261–263, 1990. 179. F Queirolo, P Valenta, S Stegen, SW Breckle. Heavy metal concentrations in oak wood growth rings from the Taunus (Federal Republic of Germany) and the Valdivia (Chile) regions. Trees 4:81–87, 1990. 180. P Tru¨by, M Lindner. Mangan-Verteilungsmuster in Fichten (Picea abies Karst.). Angew Bot 64:1–12, 1990. 181. P Tru¨by, HW Zo¨ttl. Schwermetallbelastung und Gesundheitszustand von Waldba¨umen, Kernforschungszentrum Karlsruhe-Projeht Europaeisches Forschungszentrum fu¨r Massnahmen zur Luftreinhaltung 61:257–269, 1990. 182. DA Vroblesky, TM Yanosky. Use of tree-ring chemistry to document historical ground-water contamination events. Ground Water 28:677–684, 1990. 183. W Zech, BU Schneider, H Ro¨hle. Element composition of leaves and wood of beech (Fagus sylvatica L.) on SO 2-polluted sites of the NE-Bavarian mountain. Water Air Soil Pollut 54:97–106, 1990. 184. JV Gilfrich, NL Gilfrich, EF Skelton, JP Kirkland, SB Qadri, DJ Nage. X-ray fluorescence analysis of tree rings. X-Ray Spectrometry 20:203–208, 1991. 185. RP Guyette, BE Cutter, GS Henderson. Long-term correlations between mining activity and levels of lead and cadmium in tree-rings of eastern red-cedar. J Environ Qual 20:146–150, 1991.

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186. G Kennedy, S Bergeron. Tree rings as monitors of heavy metal air pollution histories. J Radioanal Nucl Chem 151:337–343, 1991. 187. LF Ohmann, DF Grigal. Properties of soils and tree wood tissue across a lake states sulfate deposition gradient. U.S. Dept of Agriculture, North Central Forest Exp Station, Resource Bulletin NC-130, 1991. 188. K Pernestal, B Jonsson, JE Haellgren, H-K Li. Analysis of tree samples from an acidic and a limed environment by means of PIXE. Int J PIXE 1:281–296, 1991. 189. KH Ritters, LF Ohmann, DF Grigal. Woody tissue analysis using an element ratio technique (DRIS). Can J For Res 21:1270–1277, 1991. 190. J Zayed, P Andre, G Kennedy. Variabilite´ spatio-temporelle de l’alluminium chez l’epinette noire (Picea mariana). Water Air Soil Pollut 55:337–344, 1991. 191. M Ferretti, R Udisti, E Barbolani. A new approach to experimental data evaluation from tree ring analysis. Ion chromatographic determination of P, Cl and S. Fresenius J Anal Chem 343:607–611, 1992. 192. J Hagemeyer, A Lu¨lfsmann, M Perk, SW Breckle. Are there seasonal variations of trace element concentrations (Cd, Pb, Zn) in wood of Fagus trees in Germany? Vegetatio 101:55–63, 1992. 193. DA Vroblesky, TM Yanosky, FR Siegel. Increased concentrations of potassium in heartwood of trees in response to groundwater contamination. Environ Geol Water Sci 19:71–74, 1992. 194. T Wickman. Sr isotopes and elemental concentrations of tree rings—a mirror of soil acidification? Lundqua Rep 34:334–337, 1992. 195. TM Yanosky, DA Vroblesky. Relation of nickel concentrations in tree rings to groundwater contamination. Water Resources Res 28:2077–2083, 1992. 196. J Zayed, S Loranger, G Kennedy. Variations of trace element concentrations in red spruce tree seedlings. Water Air Soil Pollut 65:281–291, 1992. 197. M Ferretti, R Udisti, E Barbolani. Mineral nutrients and trace metals in tree rings of Pinus sp. Fresenius J Anal Chem 347:467–470, 1993. 198. AG Hemmann. Umweltrelevanz von δ 13 C-Klimakorrelationen an Jahrringen rezenter Kiefern (Pinus sylvestris L.). Berichte des Forschungszentrums Ju¨lich Nr. 2805, 1993. 199. Y Katayama, K Nishimura, N Kojima, Y Ishimaru, N Okada, K Minato, M Tomita, K Yoshida. Determination of trace elements in tree rings by the PIXE method. Nucl Instr Meth Phys Res 75:212–214, 1993. 200. A Kudo, T Suzuki, DC Santry, Y Mahara, S Miyahara, JP Garrec. Effectiveness of tree rings for recording the Pu history at Nagasaki, Japan. J Environ Radioactivity 21:55–63, 1993. 201. Z Lukaszewski, R Siwecki, J Opydo, W Zembrzuski. The effect of industrial pollution on copper, lead, zinc and cadmium concentration in xylem rings of resistant (Populus marilandica) and sensitive (P. balsamifera) species of poplar. Trees 7: 169–174, 1993. 202. M Wenk. Zur Struktur unterschiedlich schadstoffbelasteter Fichtenwurzeln. Beitr Forstwirtsch Landschaftso¨kol 27:117–122, 1993. 203. C Dittmar, W Zech. Dendroo¨kologische und erna¨hrungskundliche Untersuchungen an gesunden und kranken Buchen (Fagus sylvatica L.) der nordbo¨hmischen Mittelgebirge. Dendrochronologia 12:45–58, 1994.

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204. E Forget, F Courchesne, G Kennedy, J Zayed. Response of blue spruce (Picea pungens) to manganese pollution from MMT. Water Air Soil Pollut 73:319–324, 1994. 205. E Hofmann, C Lu¨dke, H Scholze, H Stephanowitz. Analytical investigations of tree rings by laser ablation ICP-MS. Fresenius J Anal Chem 350:253–259, 1994. 206. F Hofmann. Jahrringenanalysen zue retrospektiven Untersuchung von Belastungszeitreihen im Rahmen der Umweltvertra¨glichkeitsuntersuchung sowie bei der Umweltu¨berwachung von Anlagen. Norddeutsche Naturschutzakademie Berichte 7: 38–45, 1994. 207. N Momoshima, EA Bondietti. The radial distribution of 90Sr and 137Cs in trees. J Environ Radiactivity 22:93–109, 1994. 208. N Momoshima, Y Takashima, M Koike, Y Imaizumi, T Harada. Distribution and extraction behaviour of elements in annual rings of Cryptomeria japonica and Abies firma. Bunseki-Kagaku 43:891–895, 1994. 209. R Myre, C Camire. Distribution de P, K, Ca, Mg, Mn et Zn dans la tige des melezes europeen et laricin. Ann Sci Forestieres 51:121–134, 1994. 210. P Tru¨by. Zum Schwermetallhaushalt von Waldba¨umen. Freiburger Bodenkundl Abhandl 33:1–286, 1994. 211. T Aoki, N Okada, Y Katayama, T Nagatomo. Distribution and behaviour of 90Sr and 137Cs in the tree rings of two Japanese cedar samples. Proceedings of the International Workshop on Asian and Pacific Dendrochronology, 1995, pp. 216– 221. 212. B Cote, C Camire. Application of leaf, soil, and tree ring chemistry to determine the nutritional status of sugar maple on sites of different levels of decline. Water Air Soil Pollut 83:363–373, 1995. 213. JP Garrec, T Suzuki, Y Mahara, DC Santry, S Miyahara, M Sugahara, J Zheng, A Kudo. Plutonium in tree rings from France and Japan. Appl Radiat Isotopes 46: 1271–1278, 1995. 214. L Kardell. The occurrence of various heavy metals in tree rings of oak (Quercus robur L.) and pine (Pinus sylvestris L.) after traffic-rerouting and mining shutdown. The Swedish University of Agricultural Sciences Department of Environ Forestry, Report 62, 1995, pp. 1–40. 215. WC Shortle, KT Smith, R Minocha, VA Alexeyev. Similar patterns of change in stemwood calcium concentration in red spruce and Siberian fir. J Biogeography 22:467–473, 1995. 216. P Tru¨by. Distribution patterns of heavy metals in forest trees on contaminated sites in Germany. Angew Bot 69:135–139, 1995. 217. Z Li, J-L Qian, D Planas. Mercury concentration in tree rings of black spruce (Picea mariana MILL. B.S.P.) in boreal Quebec, Canada. Water Air Soil Pollut 81:163– 173, 1995. 218. SR Poulson, CP Chamberlain, AJ Friedland. Nitrogen isotope variation of tree rings as a potential indicator of environmental change. Chem Geol 125:307–315, 1995. 219. T Suzuki. Historical monitoring of heavy metal (Cd, Cu and Hg) content analysis in annual rings of sugi trees from pollution sources. Proceedings of the International Symposium on Asian and Pacific Dendrochronology, 1995, pp. 108–113.

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14 Heavy Metal Interactions in Soils and Implications for Soil Microbial Biodiversity R. Naidu, G. S. R. Krishnamurti, W. Wenzel, and M. Megharaj CSIRO Land and Water, Urrbrae, South Australia, Australia

N. S. Bolan Massey University, Palmerston North, New Zealand

1

INTRODUCTION

Heavy metals are defined as those elements (some metalloids) that have atomic density ⬎6 g cm ⫺3. Some of these elements include cadmium (8.65), cobalt (8.90), mercury (13.50), nickel (8.90), zinc (7.10), lead (11.30), and copper (11.30). Often arsenic and selenium with a specific gravity of 5.73 and 4.79, respectively, are also included among the heavy metals. Heavy metals includes both biologically ‘‘essential’’ (Co, Cu, Mn, and Zn) and ‘‘nonessential’’ (Cd, Pb, Hg, As) elements. The essential elements are required in low concentrations and hence are known as ‘‘trace elements’’ or ‘‘micro elements.’’ The nonessential heavy metals are phytotoxic and/or toxic to animals and are known as ‘‘toxic elements.’’ Consequently, heavy metals are extensively researched in the life, 401

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agricultural, and environmental sciences. Among the heavy metals As, Se, Cr, Cu, Cd, Zn, and Pb have been most commonly studied. Other heavy metals, such as mercury (Hg) and tin (Sn), are important, especially in the mining environment and around industries involving processing of ores containing these metals. The interaction of heavy metals with soil mineral components has been recently reviewed (see, for example, reviews by Naidu et al., 1995, 1996, 1997). These studies reveal that such interactions vary with both the nature and properties of soils, with the retention of metal cations generally increasing with increasing soil pH. The implications of such interactions to soil and crop quality have also been studied, although the role of soil chemical factors on remediation of metal contaminated sites is challenging and still a subject of much interest. In contrast to metal-soil mineral interactions, only limited studies have been conducted on the impact of metals on soil biodiversity, especially in relation to both bioavailability and metal-organic interactions. Where such studies have been performed, the focus has been heavy metal toxicity and bioavailability to soil microbial populations at the population level (Brookes et al., 1986; Chander and Brookes, 1991a,b; Ma˙rtensson and Witter, 1990) or at the functional level (e.g., Giller et al., 1997). While metal contaminants may change microbial community structure and species diversity, the functional ability of the community may remain unchanged. The abundance and species composition of microbial groups reflect the total ecotoxicological, effects of heavy metals and not just the results of bioavailability at any particular time. In this chapter, by following an overview of sources of metals and their interactions in soils we aim to relate soil-metal interactions to microbial biodiversity and its potential implications to soil productivity. 2

SOURCES OF METALS IN SOIL ENVIRONMENT

Heavy metals in terrestrial and aquatic environments originate from human activities and weathering of soil parent materials. Due to their low bioavailability, the metals present in ores are often not available for plant uptake and cause minimal impact to soil organisms. The natural process of pedogenesis leads to mineral breakdown and translocation of the products, as well as accessions from dust storms, volcanic eruptions, and forest fires. Often the concentrations of heavy metals released into the soil system by pedogenic processes are largely related to the origin and nature of the parent material. Typically soils rich in heavy metals include those derived from serpentine. These soils are disproportionately rich in magnesium (Mg) and heavy metals such as Ni, Cr, and cobalt (Co) and poor in calcium (Ca). Copper-, zinc- (Zn-), and Pb-rich soils have also been reported in a number of countries (Adriano, 1986). Along with these toxic metals, many incidences of poisoning have been reported in northern India, West Bengal, Bangladesh, and parts of Asia where local people have been adversely impacted by

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ingestion of arsenic (Chakroborty and Saha, 1987; Naidu and Skinner, 1999) and selenium ingested either through water or crops produced on contaminated soils. While Bangladesh and West Bengal poisoning have been linked to the presence of elevated concentrations of As in groundwater, Se toxicity in northern India has been associated with elevated levels of Se in food produced on Se-contaminated soils. The origin of Se has been linked to naturally occurring minerals (Dhillon and Dhillon, 1990). Apart from Se and As, other metals (e.g., Cr. Ni, Pb) derived via pedogenic processes have limited impact on soil biota and crop quality. Unlike pedogenic inputs, heavy metals added through anthropogenic activities have high bioavailability (Naidu et al., 1996). Anthropogenic activities, primarily associated with industrial processes, manufacturing, and the disposal of domestic and industrial refuse and waste materials (Adriano, 1986), are the major source of metal enrichment in soils. Atmospheric pollution from lead based petrol is a major issue in many developing countries where there is no provision for the introduction of unleaded fuel. Disposal of industrial waste materials has been estimated to contaminate thousands of hectares of productive agricultural land in countries throughout the world. This is a particular problem in the Asian Pacific region where land based disposal of untreated wastes is still being practiced due to the lack of regulatory guidelines (Naidu et al., 1996). Uncontrolled mining, manufacturing, and disposal of metals and metal-containing materials inevitably cause environmental pollution. While there are vast differences in the amounts of different metals produced, soil concentrations of most metals, especially Ni, Cr, and Cd, have increased over the last 60 years (Alloway, 1990). 3

FATE OF METAL IONS IN THE SOIL ENVIRONMENT

Metal ions can be retained in the soil by both sorption and precipitation reactions. The lower the metal solution concentration and the more sites available for sorption, the more likely that sorption/desorption processes will determine the soil solution concentration (Bru¨mmer et al., 1983; Tiller et al., 1984a,b). However, the fate of metals in the soil environment is dependent on both soil properties and environmental factors. Following input into the soil environment, heavy metals interact with the soil mineral and organic phases (Fig. 1). Both soil properties and soil solution composition determine the dynamic equilibrium between metals in solution and the soil solid phase. The concentration of metals in soil solution is influenced by the nature of both organic (citrate, oxalate, fulvic, dissolved organic carbon) and inorganic (H 2 PO 4⫺, NO 3⫺, Cl ⫺, and SO 42⫺) ligand ions through their influence on metal sorption processes (Shuman, 1986; Homann and Zasoski, 1987; Naidu et al., 1994a; Harter and Naidu, 1995; Naidu and Harter, 1997; Bolan et al., 1999a). Two reasons have been given for the effect of inorganic anions on the sorption of heavy metal cations (Naidu

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FIG. 1 Fate of pollutant and nutrient ions in soil ecosystem.

et al., 1994). First, inorganic anions form ion pair complexes with heavy metals and thereby reduce the sorption of heavy metals. Second, the specific sorption of ligand anions is likely to increase the negative charge on soil particles (Fig. 2) and thereby increases the sorption of heavy metals. A number of studies have examined the effect of inorganic anionic complex formation on the sorption of Cd 2⫹ by soils (Boekhold et al., 1993; Naidu et al., 1994b). Most of these studies have indicated that among the four common inorganic anions H 2 PO 4⫺, NO 3⫺, Cl ⫺, and SO 42⫺, Cl ⫺ has often been found to form a complex with Cd 2⫹ as CdCl ⫹, thereby decreasing the sorption of Cd 2⫹ onto soil particles (Naidu et al., 1994a). However, when the activity of Cd 2⫹ was corrected for complex formation the sorption curves for Cd 2⫹ in NO 3⫺ and Cl ⫺ media coincided. O’Connor et al. (1984) showed that while the presence of Cl ⫺ ions (as CaCl 2) decreased adsorption of Cd, SO 42⫺ (as CaSO 4) ions increased Cd sorption relative to comparable concentrations of ClO 4⫺ in Typic Torrifluvents, Ustollic Calciorthids, and Petrocalcic Paleustolls. All three soils were calcareous in nature, with pH values exceeding 7.5. Cd-chloro complexation was identified as the active process reducing Cd retention. They attributed the increased retention in the presence of SO 42⫺ to the low Ca ion activity available for competition with Cd due to the formation of the soluble CaSO 40 complex. In contrast to inorganic ligand ions, Haas and Horowitz (1986) found that in a few cases Cd adsorption by kaolinite, a variable-charge mineral, was enhanced by the presence of organic matter (OM) which was attributed to the formation of an adsorbed organic layer on the clay surface.

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FIG. 2 Effect of solution pH and organic ligand ions on ζ potential of an oxisol (Naidu and Harter, 1995).

Anion-induced cation sorption has been reported for a number of cations (Ryden and Syers, 1976; Bolland et al., 1977; Wann and Uehara, 1978; Shuman, 1986). Specific sorption of anions and cations onto variable charge components has often been shown to increase the surface negative charge (Table 1). Shuman (1986) observed that the specific sorption of anions, such as H 2 PO 4⫺ and SO 42⫺, increased the sorption of Zn 2⫹ by variable-charge components. Bolland et al. (1977) and Kuo and McNeal (1984) reported that H 2 PO 4⫺ sorption increases the sorption of Zn 2⫹ and Cd 2⫹ by hydrous ferric oxide. Anion-induced cation sorption depends on the variable-charge components of the soils. Naidu et al. (1994a) and Bolan et al. (1999a) have demonstrated phosphate and sulfate induced Cd sorption in strongly weathered oxisols from Australia and less weathered but net positively charged andept from New Zealand. The above studies suggest that two different mechanisms may operate depending on the nature of the soil surface (Harter and Naidu, 1995). Soils that have a high affinity for ligand anions may enhance metal ion adsorption through modifications in the soil surface charge density. In strongly weathered soils, ligand adsorption on clay surfaces enhance metal adsorption. This is illustrated by the recent studies of Naidu et al. (1994a) and Bolan et al. (1999a), who reported that the adsorption of specifically sorbing ligands such as PO 43⫺ and SO 42⫺ in strongly weathered and variable-charge soils can induce Cd

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TABLE 1 Increase in Surface Charges Due to Specific Adsorption of Anions and Cations (Bolan et al., 1999b) Soil constituents Iron hydrous oxide Allophane Soil Soil

Soil Soils Aluminum oxide Soils Soil Soil

Solute

pH

Charge added (mol mol ⫺1 anion)

Phosphate

6.5

1.25

Phosphate Phosphate Sulfate Phosphate

Phosphate Phosphate Sulfate

5.1 6.5 6.5 5.0 6.5 7.5 7.0 5.8 5.0

0.5 0.65 0.26 0.38 0.47 0.77 0.35–0.7 0.52 1.06

Sulfate Calcium Calcium

5.6 5.8 5.8

0.25 0.35–0.58 0.52

Ref. Bolan et al. (1985) Rajan et al. (1974) Bolan et al. (1986b) Sawhney (1974)

Schalscha et al. (1974) Naidu et al. (1990b) Rajan (1978) Curtin and Syers (1990) Bolan et al. (1993) Ryden and Syers (1975)

adsorption through increased negative surface charge. A similar interaction involving organic ligands was also reported by Davis and Leckie (1978) for variable-charge oxide surfaces where certain organic ligands were found to enhance the adsorption of Cu(II) or Ag(I). An alternate mechanism, which appears to be important in temperate soils, involves metal-ligand complexation in solution and subsequent reduction in cation charge, which probably reduces adsorption (Naidu and Harter, 1998). The decrease in Cd sorption onto soils in the presence of phosphate (Krishnamurti et al., 1999) was attributed to the formation of soluble Cd-phosphate complexes such as CdH 2 PO 4⫹ (log K CdH2PO4⫹ ⫽ 2.91, Martell et al., 1997). This is consistent with the concept that free Cd 2⫹ activity, rather than total dissolved Cd concentration, is a controlling factor in Cd sorption (O’Connor et al., 1984). The relative value of distribution constant, as measured by the Freundlich parameter ‘‘a’’ derived from the adsorption isotherm of the soils, had shown that phosphate-retarded Cd sorption by at least one order of magnitude. The Freundlich parameter a for the mildly alkaline (pH 7.8) soil (organic C 3.9%) was 68,071 compared to the value of 2334 for mildly acidic (pH 6.1) soil (organic C 3.0%). This was consistent with the behavior of a cationic heavy metal (Welp and Bru¨mmer, 1997). This mechanism may also be operative in high-pH (limed) tropical soils in which anion adsorption will be low and any effect on adsorption will be through ligand-metal complexation process.

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However, the nature of interaction between metals and soil colloidal particles varies with the nature of soil type and the sorbing surface. Depending on the nature of metal-soil interactions, sorption reactions may be defined as nonspecific or specific. Nonspecific adsorption from outer-sphere surface complexes is also referred to as physical adsorption (Sposito, 1984). If the physical adsorption process was the only factor controlling metal sorption, then the maximum adsorption capacity of the soils should be dictated by their cation exchange capacity (CEC). However, in many soils the amount of heavy metal sorbed exceeds the CEC of the soils. This suggests that in addition to physical adsorption other processes, such as precipitation and specific sorption, also contribute to metal retention in soils. Specific sorption refers to exchange of heavy metals with surface ligands to form covalent bonds with lattice ion. Specific adsorption results in heavy metal adsorption far greater than that expected from cation exchange. In specific adsorption, the ions penetrate the coordination shell of the structural atom and are bonded by covalent bonds via O and OH groups to the structural cations. Thus, specific adsorption of metal ion may influence the surface charge characteristics, ζ potential, and electrophoretic mobility of soil particles, demonstrated by lowering of the point of zero charge (Bolan et al., 1993). Specific adsorption is strongly pH-dependent, increasing with pH for cationic species and decreasing with pH for anionic species. However, the extent of hydrolysis is pH-dependent and therefore influenced by the soil/solution ratio. The metal ions which hydrolyze range from hard metal ions, i.e., those which include the alkali metal and alkaline earth metal ions, to moderately soft metal ions, such as Cd 2⫹, Pb 2⫹, and Hg 2⫹. Precipitation appears to be the predominant process in high pH soils and in the presence of anions such as S 2⫺, CO 32⫺, OH ⫺, and PO 43⫺ and when the concentration of the heavy metal ion is high (Naidu et al., 1997). Coprecipitation of metals, especially in the presence of iron oxyhydroxide, has also been reported and often such interactions lead to significant changes in the surface chemical properties of the substrate (Table 2). Since the product of the hydrolysis reaction is H ⫹, metal ions are generally classed as Lewis acids (after G. N. Lewis). Pearson (1963) further classified metal ions into two groups, depending on whether they formed their most stable complexes with C, N, or O, the first ligand atom from groups V, VI, and VII (hard acids, class a), or whether they formed their most stable complexes with the second or a subsequent member of each group (soft acids, class b). This classification, known as the hard and soft acid base theory (HSAB), states that hard Lewis acids prefer to react with hard Lewis bases and soft acids prefer to react with soft bases. Hard acids are usually small have a high positive charge, and low polarizability, and do not contain unshared pairs of electrons in their valence

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TABLE 2 Coprecipitation of Minerals in the Presence of Iron and Manganese Oxyhydroxides, Calcite, and Clay Minerals Mineral

Coprecipitation

Fe oxides Mn oxides Calcite Clay minerals

Ni, Cu, Zn, Mo Co, Ni, Zn, Pb Mn, Co, Cd Ni, Co, Zn, Cu, Pb

shells (Pearson, 1963). Trivalent metal ions such as Fe 3⫹ and Al 3⫹ are generally hard because of their high ionic charge. Hard bases have high electronegativity, low polarizability, and are hard to oxidize. Soft acid and soft bases have properties opposite that of the hard acid and bases. Table 3 lists the metal ions and the inorganic ligand anions according to hard and soft acid–base concept. Generally, the hard metal ions include the plant macronutrient and secondary nutrient metal cations (Ca, Mg), whereas the borderline acids include the micronutrient elements such as Mn(II), Fe(II), Co, and Zn. The hydroxy species of metal ions (MOH) is more strongly retained than the corresponding metal ions. Hence, specific adsorption increases with a decrease in the pK value for hydrolysis reaction: Cd(pK 10.1) ⬍ Ni(9.9) ⬍ Co(9.7) ⬍ Zn(9.0) ⬍⬍ Cu(7.7) ⬍ Pb(7.7) ⬍ Hg(3.4) Other chemical interactions that contribute to metal retention by colloid particles include complexation reaction between inorganic and organic ligand ions. As might be expected, the organic component of soil constituents has a high affinity for heavy metal cations because of the presence of ligands or groups that can form chelates with metals (Harter and Naidu, 1995; Naidu and Harter, 1998).

TABLE 3 Some Hard and Soft Acids and Bases Hard Acids Bases

2⫹

3⫹

3⫹

Soft 3⫹

Sn Cr , Co , Fe , As 3⫹ H 2 O, OH, O, ROH, COO ⫺, CO 32⫺, NO 3, PO 43⫺, SO 42⫺, ClO 4⫺, F ⫺

⫹

⫹

Borderline 2⫹

2⫹

Cu , Hg , Pb , Cd , Hg 2⫹ ⫺ H , S 2⫺, SH, I ⫺

2⫹

Fe , Co 2⫹, Ni 2⫹, Cu 2⫹, Zn 2⫹, Pb 2⫹ Cl ⫺, NO 2⫺, SO 32⫺, Br ⫺

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The functional groups include carboxyl, phenolic, alcoholic, and carbonyl groups (Stevenson, 1977). With increasing pH, the functional groups dissociate, thereby increasing the affinity of ligand ions for metal cations. The order of metal cations complexed by organics is as follows: Cu 2⫹ ⬎ Fe 2⫹ ⬎ Pb 2⫹ ⬎ Ni 2⫹ ⬎ Co 2⫹ ⬎ Mn 2⫹ ⬎ Zn 2⫹ However, the extent of metal–organic complex formation varies with a number of factors, including temperature, steric factors (e.g., geometry and flexibility), concentration, etc. All of these interactions are controlled by solution pH and ionic strength (IS), nature of the metal species, dominant cation, and inorganic and organic ligands present in the soil solution. Thus, metal sorption by soils depends on the chemical nature of the metal species present in the soil solution and the nature of constituent soil minerals. 4

METAL INTERACTIONS VS. SOIL SOLID PHASE COMPOSITION

Irrespective of the nature of interaction between the metals and soil colloidal particles, following sorption, metal ions redistribute among organic and mineral soil constituents. Speciation studies suggest that the majority of metals (Cd, Cu, Zn, Pb, A) are associated with organic matter and inorganic Fe and Al oxides in soils. Factors affecting the distribution of a metal among different forms include pH, ionic strength of the soil solution, the solid and solution components and their relative concentration and affinities for the metal, and time (Jones and Jarvis, 1981; McBride, 1991; Foerstner, 1991; Ritchie and Sposito, 1995). A very large number of sequential extraction schemes have been used for soils, generally attempting to identify metals held in any of the fractions: soluble, exchangeable, adsorbed/carbonate, organically bound, Mn oxide–occluded, amorphous mineral colloid–occluded, crystalline Fe oxide–occluded, and residual or lattice mineral– bound. The most commonly used schemes are based on that designed by Tessier et al. (1979). Metal fractionation using the sequential extraction techniques have primarily been used to identify the fate of the metals applied in waste sludges and in soils contaminated with smelters and mine drainage wastes (Petruzelli et al., 1981; Sposito et al., 1982; Lake et al., 1984; Chang et al., 1984; Hickey and Kittrick, 1984; Dudka and Chlopecka, 1990). These studies suggest that treating the soils with sludges or wastes shifts the solid phases of the metals away from immobile fractions to forms that are potentially more mobile, labile, and bioavailable. For example, Dudka and Chlopecka (1990) found that residual forms of Cd, Cu, and Zn decrease from 34–43% to 6–34%, while the more easily available forms increase by treating soils with sludge (Table 4). Elaborate sequential extraction schemes and geochem modelling have frequently been used to identify the distribution of different species of the metal

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TABLE 4 Solid-Phase Metal Fractions in Soils Determined Using Sequential Extraction Scheme Cd Fraction Exchangeable Carbonates Fe oxides Organic Residual

Control 0.15 0.08 0.2 0.15 0.5

(13) (7) (17) (13) (43)

Cu Sludge

0.12 0.37 0.2 0.15 0.5

(15) (25) (14) (10) (34)

Control 0.11 0.2 0.16 1.8 16.5

(1) (1) (1) (10) (34)

Zn Sludge

0.16 2.7 77.2 46.6 21.7

(0.1) (2) (56) (34) (16)

Control 2.1 11.2 18.0 29.0 25.2

(3) (16) (26) (42) (36)

Sludge 112.8 178.9 145.0 61.6 28.5

(26) (41) (33) (14) (6)

All values in mg kg ⫺1, with the % values in parentheses, calculated on the basis of total metal in soils. Sources: Data from Dudka and Chlopecka, 1990.

amongst various forms of solid species (see reviews by Ross, 1994; Krishnamurti, 2000). However, very few attempts have been made to identify the particular species of the metal that contributes to bioavailability. The mechanism which operates the phytoavailability of solution and solid-phase species can be visualized as shown in Fig. 3. Using an innovative extraction scheme (Krishnamurti et al., 1995; Krishnamurti and Naidu, 2000), which estimates the species associated to the metalorganic complexes, the importance of metal-organic complexes in the bioavailability of Cd in native soils was established based on the multiple correlation

FIG. 3 Phytoavailability of solution and solid-phase metal species.

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analysis between the data on the distribution of solid species and the bioavailable Cd (Table 5). The importance of metal-fulvic complexes in the phytoavailability of Cu and Zn was also shown using this scheme for the speciation of solid phases of Cu and Zn (Krishnamurti, unpublished). This scheme is a refinement toward better understanding of solid-phase metal species in soils (Table 6). Based on the differential FTIR spectra of the metal-organic complexes extracted by the 0.1 M sodium pyrophosphate extractant used in their speciation scheme, Krishnamurti et al. (1997a) showed that Cd in soils was apparently bonded at the COO ⫺ of the carboxyl and OH of the phenolic groups. Compared with bulk soils, solid-phase speciation of Cd (added with fertilizer P) differs substantially in phosphate fertilizer–treated rhizosphere soils (Krishnamurti et al., 1996). The amounts of Cd species associated as adsorbed and with metal-organic complexes of the rhizosphere soils were appreciably higher than those of the corresponding bulk soils. The increase was attributed to the increased amount of carbonate, a product of plant respiration, and the organic acids released as root exudates, present in soil-root interface. The formation of aqueous complexes of Cd with low molecular weight organic acids (LMWOAs) of root exudates in soil rhizosphere are expected to dominate the solution chemistry of Cd. Based on differential pulse anode stripping voltametric and cation exchange resin extraction data, the dissolved Cd in soil solutions was found to be almost completely complexed with organic matter (del Castilho et al., 1993; Sauve et al., 2000). The functional group chemistry of the organic solutes was not determined in these studies but even organic acids dominated by COOH

TABLE 5 Data on Regression Analysis Between Phytoavailable Metal and Solid-Phase Species of a Few Soils of South Australia (Krishnamurti and Naidu, 2000) Metal-fulvic complexes

Exchangeable

Simple correlation coefficients Phytoavailable Cd 0.824 (0.002) 0.735 (0.01) Cu 0.944 (⬍0.0001) Zn ⫺0.669 (0.02) 0.832 (0.002) Multiple regression analysis Phytoavailable Cd ⫽ 0.0004 ⫹ 3.5676 Exch. Cd ⫹ 2.6500 -Fulvic Cd Phytoavailable Cu ⫽ 1.2528 ⫹ 6.2022 Fulvic Cu ⫺ 4.6304 Humic Cu Phytoavailable Zn ⫽ 33.582 ⫹ 8.8092 Exch. Zn ⫺ 2.3669 Fulvic Zn

Metal-humic complexes 0.168 (ns) 0.468 (ns)

R 2 ⫽ 0.915 (0.0001) R 2 ⫽ 0.942 (0.0001) R 2 ⫽ 0.789 (0.002)

Values in parentheses are the levels of significance; ns ⫽ significant at P ⬎ 0.1.

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TABLE 6 Multistep Sequential Extraction Scheme for Solid-Phase Metal Speciation of Soils Step 1 2 3

4 5

6 7 8

Species

Reagent

Shaking time

Exchangeable 10 ml of M NH 4 NO 3 (pH 7) 4–6 h 6h Adsorbed/carbonates 25 ml of M CH 3 COONa (pH 5) 20 h Metal-organic complexes 30 ml of 0.1 M Na 4 P 2 O 7 (pH 10) 30 ml of the soil extract was brought to pH 1.0 with 6 M HCl and the suspension left overnight for the coagulation of humic complexes. The suspension was centrifuged at 12,000g for 10 min. Metals associated with fulvic complexes were determined in the supernatant. Metals associated with humic complexes were determined in the residue solubilized with 0.1 M Na 4 P 2 O 7. 30 min Easily reducible oxides 20 ml of 0.1 M NH 2 OH.HCl (pH2) 2 h (85°C) Organics 5 ml of 30% H 2 O 2,2 ml of 0.02 M HNO 3 –3 ml of 30% H 2 O 2, 2 h (85°C) 30 min 1 ml of 0.02 M HNO 3 cool; add 10 ml of 2 M NH 4 NO 3 in 20% nitric acid 4 h (dark) Amorphous colloids 0.2 M NH 4 C 2 O 4 (pH3) 30 min (95°C) Crystalline Fe oxides 0.2 M NH 4 C 2 O 4 (pH3) in 0.1 M ascorbic acid Residual Digestion with HF-HClO 4

The volume of the reagent was for 1 g (⬍2 mm) sample. After each treatment, carried out at 25°C unless otherwise stated, the extract was collected by centrifugation for 10 min at 12,000g; the residue was washed once with 10 ml deionized distilled water. Data from Krishnamurti and Naidu, 2000.

groups should facilitate complexation of Cd. Indeed, Krishnamurti et al. (1997b) observed significant solubilization of Cd from neutral to slightly acid soils with 0.1–1 mM concentrations of acetic, succinic, oxalic, and citric acids. The data suggest that the Cd release was related to the stability constant of the Cd LMWOA complex (Fig. 4). Similar increases in the solubilities of metal cations (Hue et al., 1986) and nonmetals (Bolan et al., 1994) with increasing stability constants for LMWOA have been reported. A logical solution to minimize plant uptake and to protect the quality of the food chain is to render the trace metals in the soil immobile. The phytoavailability of the different forms of the solid phase species decreases in the order: soluble ⬎ exchangeable ⬎ adsorbed ⬎ metal-organic complexes ⬎ organics ⬎ Fe, Mn oxides ⬎ residual (fixed in mineral lattices). Immobilization of heavy metals, such as Pb, Zn, and Cd, was achieved by additives (zeolites, Czupyrna et al., 1989, Vangronsveld et al., 1990, Gworek, 1992; Mn oxides, Fu et al., 1991; clay-hydroxy Al polymers, Jansen et al., 1993, Mench et al., 1994) that may not

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FIG. 4 Relationship between Cd released from a few soils of Canada (Luseland, Waitville, and Jedburgh) by selected low molecular weight organic acids (LMWOAs) (concentration 0.01 M) during the reaction period of 0.25 h and the logarithm of the stability constant of the Cd-LMWOA complexes. (Redrawn from Krishnamurti et al., 1997b.)

produce any detrimental byproduct or alter the physicochemical environment of the soils to effect the plant growth, but could fix the heavy metals in their lattice structure. Organic immobilization may also be possible. For example, CuSO 4 is widely used in milking yard to treat lameness in dairy cattle. The residual solution is flushed to the effluent pond. In the pond, 99% of the free Cu 2⫹ is immobilized as organically complexed Cu.

5

BIOLOGICAL INTERACTIONS AND TRANSITIONS

An overview on biological interactions of trace metals in soils is presented in Fig. 5. The biological components of soils, plants, and microbes control OM dynamics via release of soluble and insoluble organic compounds and subsequent transformation (Whipps and Lynch, 1983). Plants act as sources of solid OM via senescence of areal parts and roots, and deposition of, e.g., muscillage in the rhizosphere. In addition, plant roots exudate soluble low molecular weight organic compounds, including sugars, amino acids, aliphatic acids, and phenolic compounds. Both soluble and insoluble organic compounds are utilized as substrate by microbes. Microbes, in particular microbial associations in the rhizosphere, absorb and transform OM while releasing microbial exudates and debris. Overall, biological transformation of OM may result either in increased molecular size and decreased solubility (humification), or in decreased molecular size, typically associated with enhanced solubility (decomposition).

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FIG. 5 Major interactions of heavy metals with plant roots, microbes, and organic components (solid organic matter, SOM) in soils. LMW-OM and HMW-OM denote low and high molecular weight organic matter in the solid phase, respectively; massive arrows denote matter fluxes or transitions, dotted arrows influences or any other kind of interaction.

The microbial transformation of OM has various implications to the fate of trace metals in the soil environment: • Changes of the solubility and reactivity (via functional groups) of OM can alter sorption reactions of metals onto OM and change the speciation of metals in the soil solution via the formation of metal-organic complexes (Senesi, 1992). • Metals bound to solid OM can be released during decomposition. • Redox and acid–base reactions associated with microbial transformation of OM can enhance or decrease metal solubility through alteration of the redox status of metals and the surface of metal sorbents (eg. hydrous oxides of Fe and Mn), metal hydrolysis, competition of protons or hydroxyl ions for adsorption sites, or proton/redox-induced dissolution of sorbents and metal compounds. Changes in acidity and redox may also affect the stability of soluble and surface metal-organic compounds.

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In addition, plant roots and microbes interact with metals via the following major pathways: • Plant roots and microbes surfaces can physically adsorb metals (Beveridge et al., 1995). • Metals are taken up by plant roots and microbes. • Plant-microbial associations are involved in the enzymatic transformation of metal-organic compounds, including volatiles, e.g., of Hg, Se, and As (Azaizeh et al., 1997). • Root and microbial exudates can increase metal solubility via the formation of metal-organic complexes (Gerke, 1992; Mench and Martin, 1991), associated with enhanced desorption (Krishnamurti), and changed speciation in the soil solution (Knight et al., 1997b; McBride, 1989). • Plant roots and microbes can modify metal solubility and metal sorption (Adriano, 1986; Ross, 1994) via excretion of protons, carboxylic acid (Ohwaki and Sugahara, 1997), or redox-active compounds (e.g., phenolics; Marschner and Roemheld, 1994). • Plant and microbial exudates can enhance weathering and transformation of soil minerals (Hinsinger et al., 1993) and thus release trace elements from these minerals and modify their characteristics (structure, surface area, functional groups) for sorption of trace elements. • In wetland soils, some plants with an aerenchyma can support an increased redox potential in their rhizosphere that causes coprecipitation of trace elements along with and adsorption of trace elements on precipitated iron oxides (Doyle and Otte, 1997). The intensity and rate of metal interactions with biota is not evenly distributed throughout the soils. Hot spots with increased activity include the rhizosphere and are typically associated with zones with large OM content (topsoil, around plant debris, organic soil amendments). Accordingly, gradients of metal solubility, speciation, and chemical extractability are found across the rhizosphere of metal-hyperaccumulating (Bernal et al., 1994; McGrath et al., 1997) and nonaccumulating plant species (Doyle and Otte, 1997; Neng-Chang and Huai-Man, 1992; Youssef and Chino, 1989). 6

ADVERSE IMPACTS OF METALS ON SOIL BIODIVERSITY

Soils are an important habitat for a wide variety of microorganisms differing in their physiology and metabolic activity. Microorganisms are of paramount importance in decomposition of organic matter and nutrient cycling in soil. Microorganisms themselves represent one of the largest reservoirs for essential soil

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nutrients and thus are considered to be a living pool of soil organic matter. There has been considerable amount of literature available on the effects of heavy metals on soil microorganisms and their functions in soil. In this section we have attempted to provide an overview of the impact of heavy metal pollution on microbial populations and their processes in the soil. 7 7.1

HEAVY METAL TOXICITY TO MICROBIAL COMMUNITIES Abundance and Biomass

The individual populations of microorganisms (bacteria, fungi, algae, actinomycetes, etc.) respond to heavy metal pollution in soils by way of increasing or decreasing their density (numbers) and biomass. The use of viable counts as colony-forming units as a measure of microbial abundance in soil has limited value considering that only 0.1 to ⬍10% of the total bacteria can be counted using various nutrient growth media. Several studies have shown a relationship between the metal concentration and the emergence of metal-resistant microbes in soils (Olson and Thornton, 1982; Doelman and Haanstra, 1979; Hallas and Cooney, 1981; Doelman et al., 1994). Diaz-Ravina et al. (1994) reported an increased bacterial community tolerance to metals (Cd, Cu, Zn, Ni with the exception of Pb) in a sandy loam soil containing 4.4% OM as determined by the thymidine (3 H) incorporation assay. In another study involving high metalcontaminated soil, viable counts (population size) and microbial biomass reached the same level as uncontaminated control soil over a period of 420 days (Kelly et al., 1999a) even though microbial diversity was altered. A severe decrease in the density of algal and cyanobacterial populations was observed in a long-term tannery-contaminated soil containing Cr as the major contaminant (unpublished work). 7.2

Diversity

Nordgren et al. (1986) has observed the physiological groups of bacteria as more sensitive to metal pollution in a soil taken from a smelter in Sweden. Using multivariate analysis, these authors studied the impact of metals and pH on numbers of bacteria producing acids from substrates such as maltose, arabinose, cellobiose, and xylose. Hydrolysis of chitin, starch, cellulose, and xylan was also considerably affected by the metals. Only few studies have indicated phospholipid fatty acid composition (PFLA) and Biolog as sensitive indicators of changing microbial communities due to metal pollution. Addition of Zn (6000 mg kg ⫺1 soil) to soil altered the microbial community structure as evidenced by shifts in both PLFA and Biolog metabolic profiles (Kelly et al., 1999a). The Zn-induced effects on PLFA profiles included a relative decrease in the fatty acid signatures

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indicative of actinomycetes and arbuscular mycorrhizal fungi. Knight et al. (1997a), using Biolog substrate utilization patterns, have demonstrated that addition of Cu (143–157 mg kg ⫺1), Cd (3.1–4.3 mg kg ⫺1), and Zn (307–354 mg kg ⫺1) to sandy loam soils of varying pH (4.5–7.0) reduced the metabolic potential of the extracted soil microbial population. In a study where soil was evaluated for the impact of sludge-borne heavy metals on microbial communities, even 18 years after termination of sludge application a decrease in several specific populations of microorganisms as revealed by Biolog and PLFA profiles was found (Kelly et al., 1999b). A severe reduction in the density and species composition of algae and cyanobacteria were noticed in a long-term tannery waste– contaminated soil. In highly contaminated soil (total Cr, 63,000 mg kg ⫺1 soil), only one type of alga (Chlorococcum type unicellular alga) occurred, in contrast to the control soil which had a total of eight algal species including three cyanobacteria (unpublished work). 8

EFFECTS ON MICROBIAL PROCESSES

The negative impact of metals on microbial processes is important from an ecosystem perspective, and any of such effects could potentially result in a major ecological perturbance. Hence, it is most relevant to examine the effects of metals on microbial processes in combination with communities. Following are some of the examples where metals have exerted negative effects on important microbial processes. The most commonly used indicators of metal effects on microflora in soil are (a) soil respiration, (b) soil nitrification, (c) soil microbial biomass, and (d) soil enzymes. 8.1

Carbon Mineralization

Soil respiration is one of the most commonly and widely used techniques to assess microbial activity and can be measured by adding an organic substrate (substrate-induced respiration) or by using the organic substance already present in the soil (basal respiration). A 10% decrease in basal respiration was observed in an organic soil (pH 6.2) by 4.98 µmol of Cd and 5 µmol each of Cu, Zn, and Ni (Lighthart et al., 1983). The metal toxicity was more pronounced in a sandy loam soil (pH 4.9) where 0.09 µmol Cd, 1.57 µmol Cu, 0.48 µmol Pb, 0.15 µmol Zn, and 1.7 µmol Ni g ⫺1 soil decreased basal respiration by 17, 25, 25, 21, and 28%, respectively. Valsecchi et al. (1995) studied the effect of metal pollution on the soil microflora in 16 different agricultural soils containing varying amounts of organic matter (0.9–7.10%) and pH (5.7–7.0). Ammonium acetate-EDTA– extractable heavy metal concentrations in these soils were in the following range (mg kg ⫺1): Zn, 14–4884; Pb, 9–847; Cu, 8–276; Ni, 1.95–30.0; Cr, 0.05–31; Cd, 0.5–49.1. This study showed a close positive relationship between metals

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and organic carbon in soils, while a negative relationship was observed between total metals and soil respiration or microbial biomass. These results clearly indicate that heavy metals cause an alteration in soil C cycle resulting in a decrease in the net mineralization of soil organic matter and further suggests the importance of soil properties in exerting metal toxicity toward microflora. 8.2

Nitrogen Transformation

Metals have been shown to be both stimulatory and inhibitory to nitrogen mineralization in field and laboratory studies, depending on the metal dose (Baath, 1989). Nitrification is the most sensitive parameter among various steps in the nitrogen cycle. Ammonification of organic nitrogen to ammonium ions and subsequent nitrification of ammonium to nitrate were inhibited by 100 mg kg ⫺1 Zn, 100–500 mg kg ⫺1 Pb and Cr, and 10–100 mg kg ⫺1 Cd. Brooks (1984) observed no effect on nitrogen mineralization in sludge-amended soils. Mineralization of sludge-derived OM did not differ significantly among the metal-contaminated and uncontaminated soils (Johnston, 1989). However, a low turnover of organic matter was reported in metal-contaminated soils (Chander and Brooks, 1991a,b). Thus, the results obtained by different investigators on the effects of metals on nitrogen mineralization are partly contradictory (Duxbury, 1985; Baath 1989; Chander, 1991), suggesting that caution be exercised while interpreting these results. Cadmium inhibited nitrogenase in root nodules of red alder (Alnus rubra) at concentrations greater than 0.03 µmol Cd g ⫺1 soil (Wickliff et al., 1980). Interestingly, even at lower concentrations of Cd (0.09–0.9 µmol Cd g ⫺1 soil), nitrogenase activity was still affected while nodulation was stimulated. The addition of Mo and Zn stimulated N 2 fixation by Vigna unguiculata in a sandy loam soil with optimal activity occurring at 153 mol Zn ha ⫺1 and 91 mol Mn ha ⫺1 (Rhoden and Allen, 1983). The stimulatory effect of Mo and Zn might be due to their requirement as essential elements for enzymatic activity. On the other hand, Cd has no proven biological function. There have been several reports of negative effects of Cd, Cu, Co, Ni, and Zn on both nitrogen fixation and nodulation of plants (Porter and Seridan, 1981; Vesper and Weidensaul, 1978; McGrath et al., 1988). Heterotrophic nitrogen fixation was shown to be sensitive to metal pollution. For example, 50 mg kg ⫺1 Cr and 50–200 mg Cu kg ⫺1 soil (Skujins et al., 1986) and 2–4 mg Cd kg ⫺1 (Coppola et al., 1988) were reported as sensitive to heterotrophic N 2 fixation. However, Lorenz et al. (1992) opined that heterotrophic N 2 fixing bacteria are not ubiquitous or active enough to be considered as indicator organisms. Cyanobacteria are more widespread than the other free living diazotrophic microorganisms and thus are very important for nitrogen economy of soils. These

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organisms grow autotrophically on the soil surface and utilise sun light as an energy source. Brooks et al. (1986) studied the growth and nitrogen fixation by cyanobacteria on metal-polluted soils. These investigators found no or little growth and nitrogen fixation on metal contaminated sludge-amended soils even after 120 days. Ma˙rtensson and Witter (1990) also reported the absence of cyanobacterial growth on sludge treated soils where as farmyard-manure-amended soils developed algal growth. Using sludge-amended soils, Fliessbach et al. (1989) reported a large decrease in cyanobacterial counts in metal-contaminated compared to uncontaminated soils. 8.3

Soil Enzymes

The measurement of the activity of various enzymes in soil has been suggested as an index of microbial activity. The activity of dehydrogenase has been shown to be decreased in many metal polluted soils whereas phosphatase was unaffected (Brooks et al., 1984; Reddy et al., 1987). Dehydrogenase is intracellular, present only in living organisms, and represents the active biomass whereas phosphatase is extracellular. Dehydrogenase activity has been suggested as a more sensitive

FIG. 6 Flow chart of major components involved in adverse metal impacts on soil microbes.

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indicator of sludge metal pollution in soil (Fliessbach and Reber, 1990). In another study, dehydrogenase activity was severely affected by Zn at a concentration of 6538 mg kg ⫺1 soil (Kelly et al., 1999a). Kuperman and Carreiro (1997) reported considerable inhibition of a range of soil enzymes (N-acetylglucosamidase, β-glucosidase, endocellulase, acid and alkaline phosphatase) by As, Cd, Cr, Cu, Ni, Pb, and Zn concentrations ranging from 7.2 to 48.1 m mol kg ⫺1 soil. However, the decreased activities of dehydrogenase observed in Cu-polluted soils require caution in interpreting those results since dehydrogenase measurements based on triphenylformazan formation was shown to be affected by abiotic reaction with Cu (Chander and Brooks, 1991c). There are few examples showing that metal speciation and fractionation may be a key to improve our understanding of metal impacts on soil microbes. Chaudri et al. (1999) established a relation between free Zn 2⫹ activity and inhibition of microbial activity using a bisensor test. Other approaches have employed chemical fractionation of metals in soils to establish thresholds that may be used also for legislative purposes in soil protection policy. Dwelling on the above discussion we derived a flow chart showing the major components involved in adverse effects of metals on soil microbes (Fig. 6). 9

ADVERSE IMPACTS OF METALS ON FOOD QUALITY AND IMPLICATIONS TO ANIMAL AND HUMAN HEALTH ISSUES

Food quality, specifically the introduction of unwanted elements such as heavy metals into food crops, is one of the major issues concerning regulators and the public throughout the world. Besides impacting on animal and human health, unrealistically high levels of metals also constrict free trade between countries leading to significant implications to socioeconomic issues. As many of the elements or compounds in question are present in foods in trace quantities, their presence has only relatively recently been cause for concern as improvements in analytical techniques have indicated the level of contamination. Cadmium is the element of most concern as it has a relatively high transfer coefficient from soil to plant (compared with As, F, Hg, or Pb), and human dietary intakes of this element are under scrutiny in many countries. Following ingestion of food containing Cd, a large proportion of the Cd is excreted by the body. However, a small proportion is retained, mostly in the kidneys and liver, so that it may accumulate in these organs. It is thought that above a level of 200 mg Cd g ⫺1 of kidney (wet), pathological symptoms may appear (Friberg et al., 1985). However, many of the studies examining the effect of Cd on body functions have used short-term high-dose techniques, often with pure Cd salts.

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Most of the Cd accumulated in grazing animal is derived mainly from the pasture intake. Bramley (1990) estimated that annually approximately 55 and 275 mg Cd is ingested per sheep and cattle, respectively, through the intake of herbage. Work in New Zealand has indicated that cattle and sheep may ingest 1–10% and ⬎30%, respectively, of their dry matter intake in the form of soil. In areas where soil contains high Cd, e.g., as a result of sewage sludge and fertilizer application, soil ingestion is expected to play a significant role in Cd uptake by the farm animals. Roberts et al. (1994) have shown that even though sheep ingest 36–46 kg soil per year, this only contributes 5–8% of the total Cd intake for the lax- and hard-grazed flocks, respectively. Ruminants do not have a homeostatic control mechanism for regulating Cd absorption or excretion which is affected by the level of dietary Cd content. Although intestinal uptake of Cd has been estimated to account for more than 90% of the total Cd absorbed, most of the Cd ingested is discarded. About 80– 90% of the total ingested Cd is excreted in the feces and only 0.05% is excreted in the urine. Most dietary Cd is bound to metallothionein and is absorbed intact into the circulation. Over 50% of the Cd ultimately accumulates in the liver and kidney. In animals, kidney and liver Cd accounts for 50–70% of the total Cd, with kidney having a higher Cd concentration than the liver. Other organs, such as pancreas, spleen, heart, brain, and testis, together with muscle and fat, accumulate small amount of Cd. In blood Cd is associated with albumin-like protein which is transported to kidney where it is filtered through the glomerulus and reabsorbed by the proximal tubules (Friberg et al., 1985). In animals, the effects of prolonged exposure to abnormal levels of Cd include testicular necrosis, placenta destruction, abortion, teratogenic malformations, renal damage, osteomalacia, immunosuppression, pulmonary edema, and emphysema. In humans, severe Cd exposure can result in emphysema, bronchitis, ulceration of nasal mucosa, renal dysfunction, liver necrosis, and anemia, hypertension, skeletal deformities, prostate and lung cancer, and teratogenesis (Chowdhary and Chandra, 1987). Different countries have set guidelines on the maximum permissible levels for Cd in various meat products (Table 7). In an earlier survey of cattle, pigs, and sheep in New Zealand, the mean Cd concentration in kidney cortex, liver, and muscle was less than 0.4, 0.1, and 0.05 mg Cd kg ⫺1 wet weight, respectively (Solly et al., 1981). These values are well below both the concentration measured in other countries (Table 4) and the maximum permissible concentration stipulated by many countries (Table 3). More recent testing of animal offal indicated that some 22–28% of sheep and 14–20% of cattle between 1988 and 1991 had kidney Cd contents greater than permissible level of 1 mg Cd kg ⫺1 (Roberts et al., 1993). In general, older animals had higher kidney Cd contents as these animals had longer exposure to

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TABLE 7 Legal Maximum Permissible Limits for Cadmium in Muscle, Liver, and Kidney of Sheep and Cattle in Different Countries Maximum permissible limit (mg Cd kg ⫺1 fresh weight) Country The Netherlands Australia Hungary New Zealand Denmark

Muscle

Liver

Kidney

0.05 0.2 0.5 1.0 0.1

1.0 1.25 0.5 1.0 —

3.0 2.5 0.5 1.0 0.5

Cd in their environment and hence greater opportunity to consume and retain Cd. Feral deer and sheep isolated from human intervention also showed agerelated retention of Cd in kidney tissue as a result of exposure to naturally occurring Cd in the environment. The average annual consumption of red meat per New Zealander (70-kg adult) is estimated to be 80 kg. Based on a median Cd content of 0.01 mg kg ⫺1 wet weight, the annual intake of Cd is expected to be 0.8 mg, which is equivalent to a daily intake of 2.19 µg Cd. This is considerably less than the maximum safe level of 70 µg Cd day ⫺1 (Peters, 1988) from all sources in the diet. Although offal, such as liver and kidney, contains more Cd than other meat, the small quantity of offal consumed by the New Zealand population is unlikely to contribute significantly to the total Cd intake. However, in sections of population and in pets where offal forms a larger portion of the diet, Cd intake from this source could be a major concern. The ‘‘population critical concentration’’ of Cd at which 10% of the human population would exhibit signs of renal impairment is about 160 mg Cd kg ⫺1 wet weight (Friberg, 1984). From Table 8 it can be shown that it is unlikely that Cd input from meat exceeds the critical concentration in the lifetime of a New Zealander. The effect of prolonged exposure to low levels of Cd in foods (where Cd is often complexed by organic anions) has not been evaluated. To date the only reported human health problems directly attributed to Cd have been caused by occupational exposure in the mining or metallurgical industries, or by the consumption by a malnourished population of food contaminated by industrial wastes (the famous outbreak of ‘‘Itai-Itai’’ disease in Japan). Some workers argue that risks due to dietary intake of Cd are low (Ryan et al., 1982), although the World Health Organization/Food and Agriculture Organization Expert Committee on Food Additives in 1972 proposed a maximum tolerable level for intake by hu-

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TABLE 8 Cadmium Concentration in Liver and Kidney Samples of Cattle and Sheep in Various Countries Cd concentration (mg Cd kg ⫺1 fresh weight) Country

Animal

Liver

Kidney

Ref.

Australia

Cattle Sheep Cattle Sheep Cattle Cattle

0.18 0.30 0.10 0.10 0.11 0.21

0.30 0.96 0.25 0.25 0.36 0.55

Langlands et al. (1988)

New Zealand Netherlands USA

Solly et al. (1981) Vos et al. (1987) Mussman (1975)

mans of 6.7–8.3 mg Cd kg ⫺1 wk ⫺1 body weight. This was revised in 1980 to 7 mg Cd kg ⫺1 wk ⫺1 and after review recently (1993) was modified to 5 mg Cd kg ⫺1 wk ⫺1. In keeping with these regulations some countries have adopted food regulations controlling Cd concentrations.

10 CONCLUSION Based on the available studies some generalizations can be made regarding the impact of metals on soil biota and their processes. These studies raise concern about long-term effects of these metals at large concentrations in the environment since heavy metals can persist for tens of thousands of years, unlike their organic counterparts, which can be mineralized. It would seem that soil respiration, organic matter degradation, nitrification, and autotrophic nitrogen fixation are among the processes most sensitive to heavy metals, while total bacterial numbers, biomass, and enzyme activities are quite insensitive. This is because soil microbial communities can change without any measurable change in total microbial biomass, which probably represents the metal-resistant microorganisms. Many of the studies involving metal effects on soil biota did not include soil analysis data, which makes it difficult to accurately interpret the results. Hence, there is a need for inclusion of quantitative analysis of soils in long-term studies for improved understanding of the potential impact of metal contaminants on terrestrial biota. It is equally important to identify the bioavailable and thus potentially toxic metal fractions and species in contaminated soils and establish relations between simple predictors of such fractions and metal impact on microbial functions related to soil quality.

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Vesper SJ, Weidensaul TC. (1978) Effects of cadmium, copper, and zinc on nitrogenfixation by soybeans. Water Air Soil Pollut 9:413–422. Vos G, Hovens JPC, Delft WV. (1987) Arsenic, cadmium, lead and mercury in meat, livers and kidneys of cattle slaughtered in the Netherlands during 1980–1985. Food Add Contam 4:73–78. Wann SS, Uehara G. (1978a) Surface charge manipulation in constant surface potential soil colloids: I. Relation to sorbed phosphorus. Soil Sci Soc Am J 42:565–570. Welp G, Bru¨mmer GW. (1997) Microbial toxicity of Cd and Hg in different soils related to total and water-soluble contents. Ecotoxicol Environ Safety 38:200–204. Whipps JM, Lynch JM. (1983) Substrate flow and utilization in the rhizosphere of cereals. New Phytol 95:605–623. WHO (1994) Assessing Human Health Risks of Chemicals: Derivation of Guidance Values for Health-based Exposure Limits. No. 170: pp 73. Wickliff C, Evans HJ, Carter KR, Russell SA. (1980) Cadmium effects on the nitrogen fixation system of red alder. J Environ Qual 9:180–186. Youssef RA, Chino M. (1989) Root-induced changes in the rhizosphere of plants. II. Distribution of heavy metals across the rhizosphere in soils. Soil Sci Plant Nutr 35: 609–621.

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15 Behavior of Heavy Metals and Their Remediation in Metalliferous Soils Arun B. Mukherjee University of Helsinki, Helsinki, Finland

1

INTRODUCTION

Soil is a part of the ecosystem. It consists of mineral materials, plant roots, microbial and animal biomass, organic compounds, water, and a gaseous atmosphere (1). All of these are unevenly distributed; in addition, heavy metals from natural and anthropogenic sources form great sinks in the soil bed. Soil also serves as a chamber for the decomposition of organic wastes and production of greenhouse gases such as CO 2 (68). The concentration of heavy metals in the environment is decreasing as well as increasing. For example, in the developed countries, low concentration of heavy metals in soil compartments is expected to be a new problem in the 21st century. On the other hand, underdeveloped countries must deal with high levels of toxic metals in their air, water, and soil compartments. For these reasons, problems related to contamination of soils in recent years have been addressed by not only the scientific community but by decision makers and economists as well. Heavy metals in the soil bed enter from both natural and industrial sources. Natural sources include wind-borne soil particles, sea salt 433

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spray, volcanoes, forest fires, and biogenic processes, whereas anthropogenic sources include power generation, industrial processes, use of metals, and vehicular discharge of pollutants. On a global scale, the emissions of Pb, Cd, V, and Zn from industrial sources exceed those from natural sources by factors of 12, 5, 3, and 3, respectively (19). There is little information on the global flux of heavy metals stemming from the above-mentioned sources and their deposition in the soil bed. But there are some studies done in the 1980s that can still be considered as state-of-the-art work (Tables 1 and 2). In addition, the spreading of P–fertilizers, sewage sludge, lime, manure, composts, high metal content of parent rocks, trash, ashes, and slags will add further hazardous pollutants, especially Hg, Pb, Cd, Zn, and V in soils, and these are challenging problems for society (20). Many consumer goods, including lead car batteries, nickel-cadmium batteries, thermometers, and fluorescent tubes, contain substantial amounts of heavy metals that will no doubt become a threat for contamination of soils; soil covers 30% of the surface of the earth where plant species grow, and there is strong competition between man, nature, and toxic metals (4). Each soil type, however, has limited capacity to hold heavy metals, which is why there is growing concern that the anthropogenic input of heavy metals in soils in different areas of this planet may exceed the limit (21,22). Environmental preferences are specially attributed to groups of heavy metals in different compartments of the ecosystem due to their toxicity, including Cd, Pb, Zn, Ni, Cr, Hg, Cu, and As for agricultural soils; Cd, Pb, Zn, Cu, and Ni for forest soil; Cd, Pb, Zn, Hg, Cu, and As for aquatic systems; and Pb, Cd, and Hg for the atmosphere (64). A number of studies have been made on the behavior of heavy metals in different compartments of the environment. It is observed that each metal behaves in a different way in the ecosystem due to different thermodynamic and physicochemical properties. For example, Cd is deposited on soils in the particulate forms, e.g., in the solid phase, whereas Hg and As exist in the atmosphere predominantly in the vapor phase. Consequently, Hg has a tendency to be re-emitted to the air once deposited on the surface of soils (5). Deposition rate of Cd is maximal in industrial countries, but in other regions of the planet where industrialization is limited, Cd enters into soil compartments through fertilizers, mostly manure, sewage sludge, and biosolids (6). It has been suggested that anthropogenic deposition of Cd is maximum (7,8), but there is some uncertainly when considering the partitioning of emissions of Cd and other heavy metals between natural and anthropogenic sources (9). It is clear that on the global scale atmospheric deposition of heavy metals to soils is most important, whereas on a local or regional scale, other sources may be more significant from the perspective of soil pollution. To improve the physicochemical properties of arid zone soils, addition of sewage sludge precedes the release of heavy metals whose soluble fraction may vary between 0.5% and 7.0% of the total amount of heavy metal–borne sludge.

Source Wind-borne soil particles Sea salt spray Volcanoes Wild forest fires Biogenic processes Total

As

Cd

Co

Cr

Cu

Hg

Mn

Ni

Pb

Zn

0.3–5.0

0–0.4

0.6–7.5

3.6–50

0.9–15.0

0–0.1

42–400

1.8–20

0.3–7.5

3.0–35

0.2–3.1 0.2–7.5 0–0.4

0–0.1 0.1–1.5 0–0.2

0–0.1 0–1.9 0–0.6

0–1.4 0.8–29 0–0.2

0.2–6.9 0.9–18.0 0.1–7.5

— 0–2.0 0–0.1

0–1.7 4.2–80 1.2–45

0–2.6 0.9–28 0.1–4.5

0–2.8 0.5–6.0 0.1–3.8

0–0.9 0.3–19 0.3–15

0.4–7.5

0–1.7

0–1.3

0.1–2.2

0.1–6.4

0–2.7

4.1–56

0.1–1.7

0–3.4

0.4–16

0.6–11.4

4.5–83

2.2–53.8

0–4.9

51.5–582

2.9–56.8

0.9–23.5

4.0–86

1.1–23.5

0.1–3.9

Heavy Metals on Metalliferous Soils

TABLE 1 Estimated Global Natural Emission of Heavy Metals (1000 t/annum)

Note: Natural fluxes of Mo, Sb, Se, and V were excluded from this table. Source: Data from Ref. 2.

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TABLE 2 Global Industrial Emissions of Selected Heavy Metals into Soil Compartments in the 1980s (1000 t/annum) Source Wastes Chemicals Ashes Fertilizers Biomass (Peat) Atmospheric deposition Total

As

Cd

1.3–15.1 36–41 6.7–37 — 0–0.5 8.4–18.0

1.1–14.3 0.8–1.6 1.5–13 0–0.3 0–0.1 2.2–8.4

25.5–215 305–610 149–446 0–0.4 0–0.2 5.1–38

52.4–111.6

5.6–37.7

485–1310

Cr

Cu

Hg

Mn

Ni

39.2–239 395–790 93–335 0.1–0.6 0.2–2.0 14–36

0–5.1 0.6–0.8 0.4–4.8 — — 0.6–4.3

95–415 100–500 498–1655 0.1–0.8 5.2–17 7.4–46

19.4–142 6.5–32 56–279 0.2–0.6 0.2–3.5 11–37

36.2–140 195–390 45–242 0.4–2.3 0.5–2.6 202–263

542–1403

1.6–15

706–2633

93.3–494

479–1039

Pb

Se

V

Zn

0.5–13.2 0.1–0.2 4.1–60 0–0.1 0–0.4 1.3–2.6

6.5–46 0.6–2.8 11–67 0–0.1 0.1–1.7 3.2–21

218–710 310–620 112–484 0.3–1.1 0.2–3.5 49–135

6–76.5

21.4–138

689–1954

Note: Depositions of Mo and Sb in soils were excluded from this table. Source: Data from Ref. 3.

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But it is one of the major contributions of toxic metals in soils and plants (10,11). It is true that in each country there are regulations on heavy metal contamination of soils based on total metal concentrations of sewage sludge, manure, or lime. But a better picture of bioavailability can be drawn when one considers the soil solution of a metal. The relation between solution, total metal, and total metal species is a result of environmental conditions, such as pH, ionic strength, and potential ligand concentration in soils (12). In addition, one should consider the geochemical forms of heavy metals in soils that may affect their solubility and hence influence their bioavailability (13,64). Animals play an important role in the soil compartment through different mechanisms such as inoculation, grazing, litter combination, and so on. These animal populations will be affected if they live longer in the environment of toxic metals, and they may acquire exposure resistance which will help prolong their lives. The decomposition of organic matter is regulated by the predation of one animal on another. The process can also be hampered when heavy metals, fertilizers, sewage sludge, and pesticides are present, causing a concomitant increase in the number of bacterial decomposers, and as such there will be retardation of decomposition of leaf litter due to less soil-animal interaction (14). But it is surprising that microorganisms are capable of acting directly or indirectly with heavy metals through mechanisms that include oxidation, reduction, acidification, complex formation, and bioaccumulation (Fig. 1) (17,18). In the last two decades, a huge number of publications have been cited regarding contamination of soils with toxic metals, chemicals, and organic compounds from industrial and natural sources. It is now also believed that global climate change will trigger further impacts of heavy metals on soil compartments. The flux of these toxic metals will shift from one region of the planet to another, which will cause energetic impact on the quality of soils, plants, and soil animals. Polluted soils may lead to tainting of foods and drinking water, which are the basic necessities for human beings, animals, domestic livestock, and aquatic species (69). It is impossible to give a detailed discussion of soil metal remediation mechanisms. However, the purpose of this chapter is to present an overview of the behavior of heavy metals in soil compartments, their uptake by plants, and remediation of polluted soils. Remediation processes are the innovations of engineers and microbiologists. More recently, green-plant-based processes are becoming popular in many countries. In the history of civilization, it is known that life cycles of plants have profound effects on physicochemical and biological processes that occur in the environment (81). In addition, remediation costs of polluted soils have been touched on in this chapter. The chapter is based on literature searches and discussions with people from different organizations and institutions.

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FIG. 1 Microbial weather processes in the environment (redrawn) (17,18).

2

SOILS

Soils stem ultimately from rocks and are a vital compartment of the ecosystem. Rocks consist of mineral and organic solids, aqueous and gaseous compounds. Upon weathering, rocks produce the parent material and when in turn the parent material weathers, it produces soil. A number of factors are responsible for the formation of soils, including parent material, climate, life, topography, and time (15). 2.1

Parent Material

Parent materials may be cited as sandstone, shale, clay, and limestone deposited at different places of the planet. Due to the interaction of plants and rocks, new chemicals are formed. Clay minerals are the products of weathering on rocks

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that have significant effects on the physical and chemical properties of the soils (16). Soils developed from different parent materials have diverse properties, texture, and colors. For example, soils that develop from granite bedrock are generally sandy loams of low fertility. However, the type of material that makes up a soil affects the movement of water and air through soil and root penetration into the soil, as well as the looseness and workability of the soil. 2.2

Climate

The climate may have a profound impact on ingredients of soils. The changes of temperature, and snow and rainfalls, will cause expansion and contraction of soil materials and thus have an effect on soil properties. Temperature has many direct and indirect effects on soil biological activity, including physiological changes, mineral weathering rates, redox potentials, and water activity. Regarding extreme temperature, there are different opinions concerning the significance of microbiological activity at or near the freezing point in soil. There are various studies that support occurrence of subfreezing bioactivity in the Arctic environment (1). 2.3

Life

One can find lichens and mosses growing in soil environment. They grow, reproduce, and die. After their death, there is formation of organic compounds that will help to produce other plants. 2.4

Topography and Time

The topography of a country is important for the growth of forest. With less water staying on the hills after rain, plants would grow more slowly. This means that there is a lack of organic matter for the next generation of plants. When the whole countryside is hilly, steep slopes develop soil slowly, whereas the flatter areas develop soils faster. In addition, soil development can be hampered due to too much water content in soils and forests. There is a big question mark regarding the time required to form soils on the earth. In Alaska, a surface layer was evident in a glacial material after only 30 years of weathering. In Finland, soil was formed from the parent rock during the glacial period, whereas in Iowa, USA, it took about 20,000 years to develop a mature soil from wind-blown loess material (15). All of the above-mentioned factors for the formation of soil can be expressed as follows: Soil ⫽ Ᏺ(Pm, C, L, T, t)

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where Ᏺ is a function, Pm is parent material, C is climate, L is life, T is topography, and t ⫽ time. 3

HEAVY METALS IN SOIL COMPARTMENTS

The term heavy metal is generally used in connection with toxicity. In the periodic table there are about 70 elements considered to be heavy metals based on their specific gravity (e.g., should be ⬎4.5 g cm ⫺3) (23,24). However, in recent years it has been said that toxicity is based on (a) dose response, (b) speciation of a metal, (c) manner of exposure, e.g., inhalation, ingestion, skin, and (d) the individual exposed (26). As 3⫹ and As 5⫹ are toxic to human beings when the concentration in drinking water exceeds more than 0.01 mg L ⫺1. But a minute concentration of As is needed for the survival of human beings (111). A group of authors (25) classified the elements in the periodic table into three groups—A, B, and borderline—based on their relative affinity for O-, S-, and N-containing ligand. In addition to the classification systems mentioned, heavy metals have also been classified according to their chemical and biological properties. In the former case, classification is based on the concepts of chemistry, e.g., acids and bases. In the latter, it is based on the toxicity of the metal. In connection with the term heavy metal, it is also to be noted that occasionally this term is replaced by the term ‘‘trace element or metal’’ or ‘‘toxic metal.’’ However, heavy metals are a subdivision of trace elements. Table 2 indicates that soils receive large volumes of heavy metals from the different wastes generated by industries. Two important sources of heavy metals in worldwide soils are ash/slag residues from power generation and the general breakdown and weathering of commercial products (such as dry-cell batteries Ni, Cd, and Hg) on land. The metallurgical industry, mining and smelting of ores, incineration of trash and ‘‘red bag wastes,’’ metal manufacturing industries, and fossil fuel–fired power plants have caused soil pollution since the industrial revolution began. From all of these sources different heavy metals, such as Cu, Pb, Zn, Hg, Cd, Ni, Cr, and V, enter directly or indirectly into the soil compartment (16,27). It has been estimated that at the beginning of the new millennium, about 240,000 km 2 of the world’s land area will have been contaminated by metalliferous mining activities (33). Metalliferous ores contain many important metals, such as Pb, Zn, Ni, Cu, Cr, Cd, and Hg. Large amounts of different kinds of ores and slag are dumped near mining and smelting areas from which leaching of heavy metals to groundwater aquifers is to be expected. Table 3 indicates the nature of concentration of metals in soils in the vicinity of mining and smelting facilities. There is an increased demand for P–fertilizers, especially in developing countries. World consumption of P–fertilizer in 1996/97 was 31 ⫻ 10 6 tonnes P 2 O 5 (28). However, the projected demand for 2025 can be about 60 ⫻ 10 6 tonnes

Source Mine Van mine Cu-Ni smelter a Ni smelter Hg mine Nuclear weapons plant Gold mining areas Ammunition industry a

Country/region

Pb

Cd

Zn

Ni

Cu

Hg

Ref.

Wales & Cornwall Wales Finland/Harjavalta Canada/Ontario Slovenia/Idrija USA/Oak Ridge Russia/Siberia Sweden/Zakrisdal

6150 — 310 — — — — 42–24,329

3.6 72 5 — — — — —

1590 20,550 — — — — — 20–23,695

— — 322 3000 — — — —

— — 5800 — — — — 6–35,128

— — — — 800–925 7000 1.24–3.59 16–197

34 35 36,43 37 41 42 48 51

Heavy Metals on Metalliferous Soils

TABLE 3 Examples of Metal Concentration (mg kg ⫺1) in Soils Near Mining and Smelting Areas

Data are for the year 1991.

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of which the major part will be consumed by Latin America, Asia, and Africa (29). Hence Cd in soils, especially in agricultural soils, may increase dramatically in the new millennium. In addition, due to the pressure for reuse of wastes, sewage sludge, manure, lime, and organic composts, these material flows will be important sources of heavy metals in soils, especially on a local scale. The concentration ranges of heavy metals in sewage sludge, fertilizers, manure, lime, and composts are given in Table 4; these indicate that sewage sludge, P–fertilizers, and composts are significant sources of heavy metals in soils. Thus, the management policies and agronomy practices in a country will also affect the total concentration of heavy metals in soils. Varied speciations of heavy metals have been detected in soil compartments with the use of improved analytical techniques. These metals can be present in the soluble form or as exchangeable elements. They are often immobile, but are associated with soil compartments, e.g., with carbonates, iron and manganese oxides, organic matter, and parent materials. These metals or compounds generate toxicity problems as a result of their shifting from soil compartments to the soil solution, plants, and groundwaters (30,31). Aluminum hydroxides can also adsorb a number of heavy metals, and in particular soil environments they can be more important than iron oxide (7). Often, different analytical procedures generate different data for metal concentrations in soils. At the beginning of the 1990s, Cd, Pb, Co, and Se in agricultural soils used for growing wheat and maize were analyzed by AAAc-EDTA for 30 countries. It was reported that mean concentration (mg L ⫺1; n ⫽ 3664) decreases in the following order: Se(16.3) ⬎ Pb(4.25) ⬎ Co(3.29) ⬎ Cd(0.1) (32). Much information is available on the atmospheric deposition of heavy metals on the global as well as local scale in forest and agricultural soils, which can be read elsewhere. However, the mobility and bioavailability of heavy metals are generally influenced by many factors, including pH, redox potential, temperature, soil constituents, cation exchange capacity, and organic compounds. Nevertheless, microbial activity and the physicochemical parameters of soil will govern the behavior of heavy metals in soil environments. 3.1

Plant Uptake of Heavy Metals

Earlier it was mentioned that heavy metals are generally immobile in soil compartments (except in very sandy soils). However, often heavy metals, including Cd, Mo, Se, As, Pb, Zn, Cu, and Hg are taken up by plants and crops. This uptake depends on (a) the concentration and speciation of the metal in soil solution; (b) the uptake by root surface and hence transfer of the metal from root surface into the root; and (c) its translocation from root to the shoot. Figure 2 indicates the

Element As Cd Cr Cu Hg Ni Pb Se V

Sewage sludge

P–fertilizers

2–26 2–1500 20–40,600 50–3300 0.1–55 16–5300 50–3000 2–10 20–400

2–1200 0.1–170 66–245 1–300 0.01–1.2 7–38 7–225 0.5–25 2–1600

N–fertilizers 2.2–120 0.05–8.5 3.2–19 1–15 0.3–3 7–38 2–1450 — —

Limestones

Manure

Pesticides (%)

0.1–24.0 0.04–0.1 10–15 2–125 0.05 10–20 20–1250 0.08–0.1 20

3–25 0.3–0.8 5.2–55 2–60 0.09–0.2 7.8–30 6.6–15 2.4 —

22–60 — — 12–50 0.8–42 — 60 — 45

Heavy Metals on Metalliferous Soils

TABLE 4 Agricultural Sources of Selected Heavy Metals in Soils (mg kg ⫺1 DW)

Note: The authors (7) cited 28 elements of which 9 are selected for this document. Source: Data from Ref. 7.

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FIG. 2 Schematic diagram of movement of heavy metals in soil-plant systems (Reproduced with permission from Ref. 40).

relationship of heavy metals between soil and plant compartments. This would imply that the route of administration is critical in determining the effects of acute toxicity. Above the critical level, Mo and Se in soils and the aquatic environment in forage crops are quite toxic to animals. As, Pb, and Hg in crops in certain amounts are harmful to consumers, but their concentrations in soils are also potentially dangerous due to their possible direct ingestion by children, wildlife, and domestic animals. In an acidic environment, Cu, Ni, and Zn are quite toxic to plants, whereas Cr(VI) is mobile and toxic to plants in soils in temperate regions (38). 4 4.1

REMEDIATION OF POLLUTED SOILS Introduction

In the last two decades, considerable attention has been paid to the problems of soils contaminated by heavy metals and their compounds in mining and industrial areas as well as in forests and agricultural soils. All soil contamination is not from single-point sources, but there are also diffuse forms of soil contamination.

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In the latter case, a large part of the land in Australia and New Zealand can be cited as an example of Cd spreading due to application of fertilizers (46). Nowadays it is well known that contamination of soils by heavy metals is difficult to treat because heavy metals cannot be destroyed, while on the other hand soils polluted by acids can be neutralized. Heavy metals associated with organic compounds in soils are also difficult to treat. In the European Union, there are at least 120,000–150,000 contaminated sites, containing 1.0 ⫻ 10 9 m 3 of wastes and contaminated soils (44). In the United States, 250 million tonnes of hazardous wastes per year is generated. These toxic metals and organic compounds may leach into the groundwaters or they can be transferred to the food chain through crops grown in waste-amended soils; Pb can be ingested by children and animals from the contaminated soils causing Pb poisoning (45,70). There are two remediation methods: (a) clean-up techniques based on fundamental removal principles, e.g., conventional remediation methods; and (b) phytoremediation and biosolids techniques. Before selecting any technique, it is necessary to know the following (47,68): Nature of concentration of heavy metals and their compounds Physical state of the heavy metal pollutants Soil type, e.g., structure, percentage of clay and humic substances Size of site, location, and the history of polluted soils It has been reported that often it is difficult to apply remediation technology due to environmental, geographic, legal, financial, social, time-related, and technical reasons. If this is the case, then techniques of immobilization, isolation, solidification, and stabilization of contaminated soils should be the aim in order to reduce risk to the ecosystem. 4.2

Remediation Principles

Remediation techniques are based on one of several principles, e.g., ex situ and in situ. Hence, it is important to know the behavior of polluted particles and nature of soils before starting up the cleaning program. The main technologies that are available are listed in Table 5. Phytoremediation, bioremediation, as well as ex situ techniques are often used for the remediation of polluted soils. The success of remediation depends very much on size, location, and background information of the polluted soils. Clean-up techniques used will also vary from slightly polluted soils to highly contaminated soils. In addition, information is necessary on how the soils have been polluted, e.g., emissions from a facility, dumping of wastes, or dumping of tailings from mines. Time scale and the ways in which soils are polluted will give valuable information of soil texture, properties of soils, geochemical conversion processes, and the distribution of heavy metals including organic compounds.

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TABLE 5 Remediation Technologies Currently Available or Under Development for Heavy Metals–Contaminated Sites Method Ex situ Solidification

Thermal treatment Washing Leaching Size, shape, and density Soil excavation In situ Solidification Volatilization of pollutants Encapsulation Attenuation Electrokinetics Phytostabilization Phytoextraction Soil amendments

Remarks Addition of cemented agent to convert contaminated soils into hard, nonporous and nonleachable material This is applicable only to mercury contaminated soils Chelate or acid extraction Batch or pile leaching with chelates or acids Technique for removal of fine particles that have the highest metal concentration Soil removal and disposal As described above Thermal treatment Cover soils with impermeable layer Dilution with uncontaminated material Introduction of electric current so that ions move to electrodes Promotes vegetable growth to immobilize metals Removal of metals by plants Reduce bioavailability of metals with P and other soil amendments

Source: Data from Refs. 47 and 49.

It is reported that the newer remediation technologies are less expensive than current remediation methods based on excavation and transportation of contaminated soils. To select any new technologies, risk assessment is necessary. The soil excavation (e.g., ex situ) technique is generally applied to soils that are contaminated by Pb in residential areas. The remaining ex situ technologies may be applied to limited amounts of highly contaminated soils. Problems arise when larger amounts of contaminated soils are treated by ex situ techniques and these are due to the generation of hazardous wastes from the process. It is somewhat costly to handle the generated hazardous wastes. Nowadays phytoremediation or phytostabilization is used in the mining and smelting areas that require remediation. It is a long-term clean-up method that is especially popular in the United States (49,50).

Heavy Metals on Metalliferous Soils

5

447

CLEAN-UP TECHNOLOGIES (EX SITU EXTRACTION)

5.1

Extraction/Wet Classification Process

Ex situ extraction is based on unit operation, which is divided into four steps: (a) Sieving, or removal of coarse and foreign substances from the feed; (b) Mixing whereby contaminated soils are mixed intensively with inorganic acids (HCl, H 2 SO 4) or organic acids (acetic acid or citric acid) of intricate agents, such as EDTA, NTA, or a combination; (c) Separation: flotation or hydrocyclone methods used to separate cleaned soil from the extracting agent; (d) treatment whereby physical or chemical methods are applied for the removal of residue from the used extracting agent. A schematic diagram of the process is given in Figure 3. 5.1.1

Disadvantage of the Process

In this separation method, low pH and an oxidizing atmosphere must be maintained to remove heavy metals from the feed. Therefore, to avoid corrosion problems, special stainless steel equipment is necessary, which is somewhat expensive. Complexing agents are also very expensive. However, these extracting chemicals are necessary for regeneration; otherwise the cleaning process becomes complex and even more expensive. This process use is limited to sandy soils and soils containing limited amounts of clay or organic compounds (47). 5.1.2

Advantage of the Process

With this process, heavy metals as well as organic compounds can be separated from the polluted soils.

FIG. 3 Process schematic of wet classification (Redrawn from Ref. 54).

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This wet method can be applied even for the removal of Hg from contaminated soils where instead of chemicals only a high-pressure water jet is used in the mixing drum. The technique is based on differences in particle size, shape, and density, settling velocity, and surface properties. Hg has affinity for organic substances and clay particles due to which wet classification is applicable for various ionic, organic, and elemental forms of Hg. In practice, 20–40 t h ⫺1 of Hg-contaminated soils can be handled by this method (52,53). Hg emission problems may occur prior to the main processes. These generally occur at the unloading station, which should be enclosed to avoid personal and occupational health hazards. 5.2

Thermal Treatment of Hg-Contaminated Soils

Hg and its compounds are volatile at low temperature (356.58°C). Most thermal soil plants are designed to remove all kind of pollutants from the soil, including petrol, oil, polyaromatic hydrocarbons, halogenated organic compounds, PCB, furane, dioxin, and so on. To remove Hg and its compounds, the facility works at about 600°C. A simplified process diagram is shown in Fig. 4. At first the soil is warmed at 100°C in a drying drum; this is followed by a heating drum where the temperature is maintained at about 600°C. Here the Hg is released and passed through different units before the exhaust gas is vented into the atmosphere. The cleaned soil is cooled and moistened (54). Different thermal techniques, such as thermal and vacuum-fired rotary dryer, and wet classification with thermal

FIG. 4 Thermal treatment for removal of mercury by engineering methods.

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treatment, for the removal of Hg have been addressed in the literature. Similarly the gas cleaning system may be different from that shown in Fig. 4. For example, the spray tower may be replaced by an activated carbon filter (47). However, there are advantages and disadvantages in the thermal treatment (54). 5.2.1

Advantage The process reduces maximum Hg concentration in treated soils.

5.2.2

Disadvantages There are problems with porous fine materials in soils, e.g., bricks and sludge. It is difficult to obtain permission from the regulatory authority for its use. There are social problems in the neighborhood. It is more expensive than any other applied technique.

5.3

Electrodialytic Remediation Method

Electrodialytic remediation is a new method for removing heavy metals from contaminated soils. Here a DC current is used as the cleaning agent together with a combination of ion exchange membranes. This method was developed at the Technical University of Denmark where electrokinetic remediation method is combined with another technique known as electrodialysis. The principal of the technique is shown in Fig. 5. Membranes 2 and 3 are in contact with the soil on one side of the membrane and with an aqueous solution on the other. Membranes 1 and 4 are used to better control the electrode reaction. After remediation, the heavy metals deposit in compartments II (cations) and IV (anions). The process can be applied for the removal of Cu, Cr, Hg, Pb, and Zn from polluted soils. When considering the electrodialytic remediation method, the following parameters should be taken into account (55). pH is crucial for the desorption of heavy metals from the soil; for Cu, Pb, and Zn, pH is maintained between 3 and 4, whereas for Cr it is about 2.5 for desorption. The current density must be kept below the limiting current density to avoid production of hydroxide ions. The hydroxide ions may stop the migration of heavy metals. Soil composition and speciation of metal play an important roll in this technology. Electroreclamation is only suitable for the in situ removal of heavy metals from the clay soils. However, few commercial applications have so far been reported for the removal of heavy metals from polluted soils by this method (47).

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FIG. 5 Principle of electrodialytic soil remediation. (Redrawn with permission from Ref. 55.)

6

CLEAN-UP TECHNOLOGIES (IN SITU EXTRACTION)

The basic principle of in situ extraction is percolation of an aqueous extracting agent through the contaminated soils in which heavy metals are dissolved during percolation. With the help of special pumping equipment, the percolate is pumped up and then treated. The extracting agent can be reused after purification. The percolation process is continued so long as removal of heavy metals from the site is satisfactory. When considering the in situ extraction, special attention should be given to the following parameters: geohydrological properties of the contaminated site, origin of the pollution, solubility of pollutants, purification process of treated chemicals, and reconditioning. Only clay and sandy soils must be cleaned up by the in situ extraction process. The success of the process depends on the source of heavy metals. If it is originally dispersed via the water phase, then there is maximal chance that in situ extraction will be successful. In addition, the heavy metals should not be strongly bound to the soil particles. If necessary, addition of such chemicals as inorganic or organic acids, complexing and oxidizing agents, and detergents may be carried out to improve extraction. Application of this process is limited. Figure 6 indicates the simplified process scheme of the in situ extraction method (47,56). There are also simpler methods for soil remediation, which are summarized in Table 6. All of these methods are readily available but expensive. A few are mentioned below:

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FIG. 6 In situ extraction (modified) (47,56).

Addition of limestone to raise the soil pH to 6.5 when soils are contaminated with heavy metals, especially Cd due to application of sewage sludge (71). Addition of clay minerals, such as vermiculite, to contaminated soils for immobilization of heavy metals and radionuclides, e.g., 137 Cs and 90 Sr (72).

TABLE 6 Selected Alleviation Methods That Are Adopted for MetalContaminated Soils Target pollutants

Soil processes involved

Limestone

Metals, radionuclides

Precipitation, sorption

Zeolite

Metals, radionuclides Metals

Fixation, ion exchange Sorption, precipitation Fixation, ion exchange, sorption

Technology

Apatite Clay mineral

Metals, radionuclides

Source: Data from Ref. 68.

Restrictions Ineffective for oxyanions; certain crops (lettuce, spinach, tuber, etc.) Insufficient data Insufficient data Clay type; short term

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Addition of organic compounds to help immobilization of heavy metals as well as to supply energy to microbe communities for decomposition and transformation of organics and metalloids (73). Addition of synthetic zeolite to contaminated soils to hinder the uptake of Cd and Pb in plants. This has been demonstrated in laboratory experiments in Poland (74). Besides the above-mentioned methods, it is a well-established fact that fertilizers are stable compounds that can bind Pb in soils. This process is quite promising in the in situ immobilization of Pb (75,76). Details may be found elsewhere (68). It is also cited that heavy metals can be stabilized by clinoptilite, hydroxyapatite, and Fe-oxide wastes in contaminated soils (77). 7 7.1

PHYTOREMEDIATION Introduction

In recent years, new technology based on the use of plants for remediation of soils, known as ‘‘phytoremediation,’’ ‘‘green remediation,’’ ‘‘botanical bioremediation,’’ and ‘‘phytoextraction,’’ has come into use. This is cost-effective in comparison with current expensive engineering methods (57–60, 102). The technology is practiced not only in the United States but in many other countries. It is an environmentally friendly green technology that nowadays is applied to both inorganic and organic pollutants present in the ecosystem, e.g., soil, water, and air (65). It is also reported that metal-containing ashes from the process can be recycled for the production of valued metals. The concentration range of a number of heavy metals in soils and their regulatory limits in the United States are shown in Table 7. However, the phytoremediation of heavy metals can be divided into three sections: (a) phytoextraction; (b) rhizofiltration; and (c) phytostabilization. It has been estimated that in 2000 the phytoremedation of heavy metal–contaminated soils in North America and Europe may be a $400 million per year business (60). In this chapter, phytoextraction and phytostabilization will be discussed in detail. Because rhizofiltration is a process for removal of heavy metals from polluted wastewater streams (62), not from soil, it will not be covered here. 7.2

Phytoextraction

Phytoextraction is generally defined as the extraction of metals by plants from soils into their roots and above-ground shoots. The system uses for the synergistic relationships among plants microorganisms, water, and soil that have evolved naturally in the terrestrial and aquatic environments over millions of years. Generally plants contribute their inherent enzymatic and uptake processes to convert

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TABLE 7 Recently Observed Concentration Range of a Number of Important Heavy Metals in Soil and Their Regulatory Limits Elements Pb Cd As Cr Hg Cu Zn

Concentrate range (µg kg ⫺1) (66)

Regulatory limit (mg kg ⫺1) a

1000–6,900,000 100–345,000 100–102,000 5.1–3,950,000 0.1–1,800,000 30–550,000 150–500,000

600 100 20 100 270 600 1500

a

Nonresidential direct soil clean-up criteria. In Clean-up Standards for Contaminated Sites, New Jersey Department of Environmental Protection (1996).

contaminants to neutral compounds. Plants also act as hosts for aerobic and anaerobic microorganisms. Their roots and shoots increase microbial activity in the direct environment by providing additional colonizable surface area, increasing readily degradable carbon substrates by discharging organics and leachates. Physically, plants decrease the movement of contaminants in soils by reducing run-off and increasing adsorption of compounds to the roots and shoots. Once the phytoremediation system starts acting in the environment, its biological components are naturally self-sustaining, powered by photosynthesis. The system can be used as a clean-up technology because it is self-sustaining, cost-effective, and aesthetically pleasing. It is now a well-established fact that by growing plants over a number of years it is possible to clean up pollutants from contaminated soils or to alter the chemical and physical nature of soil pollutants so that they do not create a risk to human health and the environment (81). Low pH of the soil plays an important factor in controlling the solubility of metals in soils. Many studies suggest that lowering the pH of a soil will induce desorption of heavy metals and thus increase their concentration in soil solutions. Increased uptake of heavy metals may be possible by maintaining lower soil pH via the addition of soil-acidifying amendments or ammonium-containing fertilizers (77). Many different kinds of plants are used to clean contaminated sites. Wild plants can also be used, but unfortunately their growth rate is slow. The aim should be to select rapid growing plants so that they have high biomass in the field. On the other hand, some authors recommend the use of hyperaccumulator species (58). It has been reported that several high-biomass crop species, related

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to the wild mustard family, may be used for the removal of heavy metals from soils. Brassica juncea (Indian mustard) has been used to collect Pb in its shoots, and the spice possesses the ability to accumulate and tolerate Cd, Cr 6⫹, Ni, Zn, and Cu (61). Based on hydroponic cultural system studies, it is reported that B. juncea and T. caerulescens accumulate heavy metals in their shoots. The latter species can accumulate higher levels of heavy metals, particularly Zn and Ni, in the shoots than the former. But B. juncea grows faster, which means it has more biomass that will collect more metals from soil compartments (60). Ebbis et al. (67) studied in the laboratory the ability of 30 plants to accumulate heavy metals from soils moderately contaminated with Cd, Cu, and Zn. These laboratory studies indicated that Brassica spp. more effective in removing Zn from soil than T. caerulescens. The reason is that the former produced 10 times more biomass. There are diverse opinions among authors regarding the clean-up of Pb by phytoremediation technology. Pb is a complex metal to remove from the soil bed because it is a soft Lews acid that bonds strongly to both organic and inorganic ligands in soils. There are three main factors that hinder the phytoextraction of Pb, including: (a) poor plant uptake due to low solubility of Pb in soils; (b) ineffective translocation from Pb in roots to shoots of a plant; and (c) toxicity of Pb to the plant tissue. Many authors cited such methods as smelting, lowtemperature ashing, incineration, and microbial degradation to handle Pb in plant tissues, but these are mostly at the theoretical stage. As of today, no Pb-contaminated site has been fully cleaned using phytoextraction technology (110). Much attention has been paid in recent years to the behavior of Cr(III) and Cr(VI) in Cr-contaminated soils. In Hudson County, New Jersey (USA), 2 million tonnes of Cr slags has been used as landfill, whereas in Finland more that 600,000 t of Cr slags has been stored near the seashore in the north. On a worldwide basis, the dumping of commercial wastes containing Cr may be the largest contributor, accounting for 51% of the total release into soils. These disposal practices are normal and create toxic problems due to the complex chemistry of Cr. Oxidation and reduction reactions can convert nontoxic Cr(III) to toxic Cr(VI) and vice versa (3,82,83). In recent years, deep-rooted plants have been used for phytostabilization of toxic Cr (58). It is also necessary to understand the role of chemicals and metal salt additions in the phytoavailability and bioavailability of metals added to soils. It is accepted that Fe and Mn oxides present in biosolids can increase the uptake of specific metals from the contaminated soils, but addition of metal salts does not have this effect (Fig. 7). A recent study suggests that ethylenediaminetetraacetic acid (EDTA) is an effective chelating compound for mobilizing the metals in the soil. Hyperaccumulation of Cd and Zn was noted by T. caerulescens (Prayon, Belgium) in the presence of EDTA (Table 8), whereas the concept of chelateassisted phytoextraction may be applied to Pb, Cu, Ni, and Zn. The authors (65) observed 28% Pb extraction from a Pb-contaminated site by using B. juncea

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FIG. 7 The model addressed increasing Cd concentration in response to increasing total soil Cd concentration. (A) Addition of a soluble Cd salt. (B) From addition of a soluble Cd salt with higher Zn concentration as a soluble Zn salt. (C) Addition of biosolids. (Reproduced with permission from Ref. 79.)

(Indian mustard plant) after application of the chelating agent EDTA. It is also reported that with the help of chelate-assisted phytoextration methods, it is possible to remove 180–530 kg Pb ha ⫺1 yr ⫺1. The volume of contaminated biomass can be reduced by ashing or composting; the residue should be treated as a hazardous waste or can be used as raw material for the extraction of metal (if economical) (65,84). 7.2.1

Phytovolatilization of Arsenic, Mercury, and Selenium

Phytovolatilization can be applied to selected metals, including As, Hg, and Se. It is not clear if terrestrial plants can volatilize As in significant quantities. As accumulates in the roots and small amounts are transported to the shoots. However, volatilization of As may occur by rhizopheric bacteria in the terrestrial plant (65). Mercury toxicity, atmospheric cycling, deposition, and accumulation as methyl-Hg in fish, birds, and mammals is well documented. More recently, it has been reported that a modified bacterial mercuric ion reductase (MerA) has been introduced into transgenic A. thaliana, which converts Hg 2⫹ into elemental

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TABLE 8 Concentration and Uptake of Zn, Cd, and Cu by T. caerulescens After Addition of EDTA (Mean Values ⫾ SD) EDTA addition ⫺ ⫹

Plant concentration (mg L ⫺1)

Plant uptake (mg pot ⫺1)

DM (g)

Zn (%)

Cd

Cu

Zn

Cd

Cu

1.50 ⫾ 0.17 1.47 ⫾ 0.19

0.58 ⫾ 0.20 1.07 ⫾ 0.54

191 ⫾ 21 167 ⫾ 54

7.4 ⫾ 0.8 8.6 ⫾ 1.0

8765 ⫾ 3190 15196 ⫾ 5839

286 ⫾ 61 253 ⫾ 116

11 ⫾ 1.9 12 ⫾ 2.5

Source: Data from Ref. 80.

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Hg (Hg 0). It is also added that these transgenic plants (Arabidopsis) are very effective at volatilizing Hg (85,98). This may be confirmed from recent research in which authors (95) used a molecular genetic approach to transfer Hg resistance genes to a variety of plant species as a potential means of phytovolatilization of Hg 0. Figure 8 indicates how the authors (95) transferred bacterial genes to a number of plants using standard methods for transgenic plant development. The same authors claimed that this method has potential to remove Hg from the contaminated soils. They suggested that it is also possible to manipulate Hg 0 as Hg 2⫹ in plant shoot tissues, which could be an additional option for Hg phytoremediation. Plants and soil microbes facilitate biosynthesis and emission of Se in the form of dimethylselenide or dimethyldiselenide. It was reported that both Se nonaccumulator and accumulator species may volatilize Se. Commercial vegetable crops (broccoli, cabbage, rice, and other plants) are quite effective in phytovolatilization, but the presence of sulfate and salinity in soils may inhibit volatilization mechanism (86,87). However, more studies are necessary to understand plant and soil microbial phytovolatilization mechanisms. 7.2.2

Removal of Heavy Metals by Hyperaccumulator Plants

The ability of plants that grow in metal-polluted soils to take large amounts of metal into their roots and shoots has been known for at least a century (88). Some plants can take up unusually large amounts of metals (10–100 times higher than normal crops) into their shoots. These plants are known as hyperaccumulators (89). Many authors have discussed specific species having remarkable hyperaccumulation capacities; these are summarized below (57,58,65,78,90–92). The hyperaccumulator species generally belong to the Brassicaceae and Fabaceae families. At least 397 species have been identified, and this number will increase based on studies of metal-enriched environments. The ecological role of metal hyperaccumulation is still not fully understood. Although different authors have different opinions, salient hypotheses include the following: (a) Hyperaccumulation is possible when plants have hypertolerance properties, which may be result of chelation and vacuolar compartmentalization. (b) Generally, the root of a plant contains more metal than the shoot. However, in hyperaccumulator plants the roots must transfer more metals to the shoots. It has been reported that in normal plant root, Zn, Cd, or Ni concentrations are 10 or more times higher than shoot concentrations, but in hyperaccumulators, shoot metal

FIG. 8 Enzymatic reactions encoded by bacterial merA and merB genes with product, (Hg 0) (95).

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concentrations (e.g., 100 mg kg ⫺1 Cd, 1000 mg kg ⫺1 Ni, or 10,000 mg kg ⫺1 Zn) may exceed root levels. (c) Hyperaccumulators can resist drought and also can interfere with neighboring plants. (d) It is also suggested that hyperaccumulators provide protection against fungal and insect attack. Recent studies suggest that Ni hyperaccumulation has a protective function against fungal and bacterial pathogens in Streptanthus polygaloides, and insect herbivore in S. polygaloides and T. montanum. An antiherbivory effect of Zn has been found in the Zn hyperaccumulator T. caerulescens (93,94). The authors (91) analyzed 168 species of Alyssum Linnaeus to identify the hyperaccumulators (⬎1000 µg g ⫺1 DM) of Ni. An additional 31 hyperaccumulators (all in section Odontarrhena) were discovered in addition to the 14 European species. Details of these species can be found elsewhere (91). There is no doubt that hyperaccumulation plants are useful in soil clean-up as they can remove significant amounts of heavy metals from contaminated soils. The efficiency very much depends on the ability of root systems of the plant to quickly transfer metals to aerial plant parts. Soils contaminated by mining and smelting are quite complex in nature. The phytoextraction will depend on root systems in the contaminated soil and the nature of the contaminated zones (96). Chaney and co-workers have done pioneering work on hyperaccumulation of metals from contaminated soils. Field results are given in Table 9. Their study indicated that soil acidification increased the uptake of Zn and Cd and the plant T. Caerulescens grew well when plant competition was limited by weed control. However, in their study, low uptake of Pb was noted which is explained by the fact that Pb hyperaccumulation is not possible when soil is treated with fertilizers to improve biomass yield. Some scientists added the chelating agent EDTA to dissolve soil Pb so that Pb-EDTA with water is up-taken by attenuated roots. Transpiration carries the Pb-EDTA from the soil into plant shoots. But the growth of plants are seized when high level of Pb is accumulated by shoots and roots

TABLE 9 Effects of Fertilizer Addition on Metal Uptake by Shoots of T. caerulescens and Lettuce on Revival field, St Paul, MN, USA, 1993

Treatment S

N

Soil pH

0 0 1 1

NH 4 NO 3 NH 4 NO 3

7.4 7.5 6.7 6.8

Source: Data from Ref. 57.

Metals in Thlaspi

Metals in lettuce

Cd Zn Pb (mg kg ⫺1 dry wt)

Cd Zn Pb (mg kg ⫺1 dry wt)

9.6 9.4 11.7 8.0

1360 1260 3100 2060

0.5 4.6 1.9 1.5

5.3 4.5 7.8 7.5

58 64 86 77

0.8 0.8 2.1 1.7

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of the plant (84). It has been stated (107) that the application of EDTA is quite expensive and not practical. However, Pb is ubiquitous in soil compartments. Today no suitable wild or crop plants have been identified for cost-effective phytoremediation, but rhizofiltration of Pb from effluents has been identified. These plants can bind Pb in their roots so that it does not leach to soils (97). But Prof. Robert Tucker of Rutgers University claimed that the ‘‘phytoremediation technique seems particularly attractive for the clean up of Pb in soils’’ (109). It has also been reported that hyperaccumulation can be achieved by genetic engineering. Genes deciphering the Cd-binding protein metallothionein have been examined in plants with successful results in terms of increased Cd resistance (99,100). 8

PHYTOSTABILIZATION

In the phytostabilization process, heavy metal–tolerant plants are used to reduce the mobility of metals so that there is less risk of heavy metals to the surrounding environment (63). These specific plant species absorb heavy metals and organic contaminants into their roots, thus reducing the mobility of the heavy metals into the groundwater as well as their availability to the food chain. This technique can be used to replace an idle soil cover at a site where natural vegetation is lacking due to the presence of high concentrations of heavy metals or physical disturbances of surface materials. In the United Kingdom, it has been reported that by using fertilizers and planting metal-tolerant plants, excellent vegetative cover was developed over metalliferous mine wastes (60). Based on this study (60), three different types of grasses were made commercially available; Agrostis tenuis cv. Goginan for acid Pb/Zn waste; Festuca rubra cv. Merlin for calcareous Pb/Zn waste; and Agrostis tenuis cv. Parys for Cu waste. Much work is going on regarding stabilization of Cd- and Zn-contaminated soils with metal-tolerant grasses in the United States (60). Recent studies also suggest that some plants may reduce metal leaching by converting metals from the soluble oxidation state to an insoluble oxidation state. It has been cited that concerning Cr, roots of the plant B. juncea can reduce available toxic Cr 6⫹ to unavailable less toxic Cr 3⫹ (101). The point is that phytostabilizing plant species must tolerate high levels of heavy metals and that they should be fixed into the roots not the shoots, thus avoiding mobilization of metals in soils as well as reducing the amount of hazardous waste biomass. Chaney and his group (79) developed special biosolids for the effective revegetation and soil restoration of such metalliferous soils. These biosolids contain fertilizers, composts, lime, and high-Fe biosolids. Tailor-made biosolids were applied to barren soils containing high concentration of Zn (1%), Cd (100 mg kg ⫺1), and Pb (100–30,000 mg kg ⫺1). By controlling the proportion of ingredients, the group (79) observed very satisfactory results on the bioavailability of Cd, Zn, and Pb.

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TABLE 10 Characteristics of the Soils and Materials at Banker Hill Site

Typical soil Soil in plots High-N biosolids Low-N biosolids Wood ash Logyard waste

Zn (mg/kg)

Cd (mg/kg)

50 10,000 620 1700 410 86

0.5 17 2.5 22 3.4 0.5

Pb (mg/kg) 10 3000 75 380 55 15

pH

C (%)

N (%)

5.5 6.4 8.0 6.5 11 7.2

3.5 0.4 32 26 18 9

0.2 0.02 5.0 2.8 0.1 0.2

Source: Data from Ref. 106.

They also pointed out that all biosolids do not contain high Fe levels or high CaO equivalent. But many industrial, urban, and agricultural products can offer high concentration of Fe and CaO equivalent for the preparation of remediation products. Composts have a special place in the application in amending soils because they can kill pathogens in biosolids and do not affect the safety of children (79). It has been demonstrated that using this type of biosolid it was possible

FIG. 9 A mixture of biosolids and ash is successful material for revegetation and erosion control (106).

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to remediate and revegetate permanently contaminated sites. Examples are Palmerton, Pennsylvania, USA (103) and Katowice, Poland (104) where mine wastes and smelter discharges damaged the ecosystem. After the application of biosolids, heavy metals and other contaminants are present at safe levels for domestic livestock and animals. Sometimes, noncomposted mixtures of biosolids, wood ash, and log yard debris are quite effective in the recovery metalliferous soils; this has been cited for soils in Bunker Hill, Idaho, USA (105). At the Bunker Hill site, mining and smelting of Zn, Pb, Cd, and As for 65 years had contaminated soils. Soils have been restored by application of biosolides (Table 10). In addition, many ameliorates are known, such as zeolites, apatites, and glauconite, that decrease the mobility of metals and their uptake by plant species through metal hydrolysis reactions and coprecipitation with carbonate (7). Before applying any biosolids it is necessary to study the efficiency of selected ameliorates as metal binders or stabilizers in contaminated soils. However, it has been observed that biosolids control metal toxicity and provide organic matter that both improves soil cultivation and acts as a substrate for soil microbes (Fig. 9). 9

REMEDIATION COSTS

Nowadays problems with contaminated soils are immense. Clean-up technologies have been developed and are cited in this chapter; however, unfortunately, due to financial, technological, environmental, geographic, and social reasons, cleaning missions are not always successful. Remediation costs based on engineering methods cited in this study are estimated to be U.S.$50–200 per tonne of soil (47). This cost was estimated at the beginning of the 1990s. Recent studies suggest that remediation of contamination soil by conventional engineering techniques may cost U.S.$50–500 per tonne of soil. Any clean-up costs must be based on soil type, physical state of pollutants, and size and location of the polluted site. Special clean-up costs may be more than U.S.$1000 per tonne of soil (81), and a group of scientists (97) estimated clean-up costs for highly contaminated soils by heavy metals at U.S.$1–3 million per acre-foot of soil (soil removal, disposal, and replacement with unpolluted soil). In Finland, a secondary Pb smelter polluted soils in its vicinity and the neighboring gardens with Pb. In the mid-1980s, smelting activities were stopped, and remediation costs by traditional engineering methods for 60,000 tonnes (35,000 m 3) of Pb-contaminated soils have been estimated to be 20–40 million marks (U.S.$ ⫽ 5.84 mk). This means that the clean-up costs per tonne of soil was U.S.$57–115. On the other hand, there are claims that phytoremediation or phytostabilization is less expensive than conventional engineering approaches, such as washing or incineration of soils. Chaney and his group (97) estimated commercial phytoremediation costs for a crop with 20 tonnes biomass per hectare per harvest containing 2.5% Zn as follows:

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500 kg/ha Zn @ $1.92/kg Zn 15 kg/ha Cd @ 7.27/kg Cd Total

⫽ $ 960/ha ⫽ $ 109/ha ⫽ $ 1,069/ha (1 ha ⫽ 100 ⫻ 100 m 2)

This means that the cost of phytoremediation per square meter is about $1.07 (for two metals) These costs vary from metal to metal. In the case of Co and Ni, it is possible to make a profit above cleaning costs by burning and recovering the metal from the biomass ash. But for Pb, U, and Se, phytoremediation costs are higher than those of Zn and Cd. Chaney and co-workers have estimated the costs of the use of biosolids and their conclusion is that it is costlier than phytoremediation techniques. In their recent experience at Bunker Hill, Idaho, remediation costs for the application of tailor-made biosolids was about U.S.$4000 per acre (4046 m 2), e.g., U.S.$0.99 m ⫺2 of polluted soil. For phytoremediation, a general overall cost would be about 50 cents m ⫺2 (pers. commun., R. Chaney, 2000). Another study estimates costs of U.S.$80 per cubic yard of soil (about 1 tonne of soils) for the phytoremediation process (109). The same author cited disposal amounts of polluted soils to be huge, e.g., an area of 2.5 acres (1.01 ha) and depth 18 in. (46 cm) generates more than 5000 tonnes of soil. In contrast, plants used to take up the metals from this same amount of soil would have only 25–30 tonnes of biomass ash after burning. The smaller ash volume from the plants is more economical to handle that the amount of polluted soils generated by conventional methods. However, there is not much information on costs for phytoremediation techniques; therefore, it is very difficult to reach a definite conclusion. 10 DISCUSSION AND CONCLUSIONS In this chapter, I have preliminary discussed sources of heavy metals in the soil compartments, their behavior in soils, uptake by plants, and remediation of polluted soils by traditional technologies with special attention to green remediation known as phytoremediation. This information may guide the policymakers in the selection of cost-effective methods for remediation of contaminated soils by heavy metals and organic compounds. Moreover, the integrated load of heavy metals in soil compartments may provide the basis for the development of policy recommendations for the long-term sustainable distribution of heavy metals through fertilizers and other components in soils. Heavy metals have been recognized as toxic chemicals causing adverse effects in the ecosystem. However, soils have been perceived as sinks for heavy metals released from anthropogenic and natural sources. The contamination of soils by heavy metals started from the beginning of the industrial revolution in the 20th century. Mining and smelting, power generation, the cement industry, chloralkali plants, secondary metal industry, and use of fertilizers and pesticides

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all place heavy metals into the soils directly or indirectly. It has been estimated that by the beginning of the 21st century, 240,000 km 2 of the earth’s land area will be contaminated by metalliferous mining activities. The number of contaminated sites is growing in the European Union and, for example, in The Netherlands numbers around 100,000, whereas in Finland it is more than 18,100. Soil consists of different minerals; clays; oxides of Al, Fe, and Mn; organic and other compounds. Heavy metals binding capacity depends on the soil composition, soil pH, and redox conditions. Organic matter is important in the immobilization of Pb. However, heavy metals behavior in the soil compartments depends on several factors, including the geochemical characteristics of the metal, plants, soil chemical equilibria, climate, and agricultural management (64). As the problem of contamination of soils is increasing, it is necessary to tackle the situation. A large number of remediation techniques have been developed over the last decade and a half. These remediation methods are based on (a) traditional clean-up technologies and (b) phytoremediation methods. Before selection of any method it is necessary to study (a) distribution and concentration of heavy metals in soils, (b) physical state of the pollutants, (c) soil properties, e.g., sand and clay/loam content values, particle size distribution and percentage of organic compounds, humic substances in soils, and (d) background information of the site, e.g., how the site has been contaminated and for how long it has been contaminated. Traditional clean-up technologies are based on incineration of soils, adsorption and desorption processes, solubility, chemical treatment, and biodegradation. For cleaning soils by traditional clean-up technologies, then, one or more methods are generally applied. However, the experiences gained with these processes indicates that they are not capable of solving all problems. In the Western world, many traditional clean-up methods have been developed. Each method has its own character based on the type of soil and the physical properties of the pollutants, respectively (108). The total number of contaminated sites found in Europe or the United States is not clear, but it is clear that the number of sites is increasing exponentially. The costs of dealing with contaminated soils are enormous. In The Netherlands, the estimated cost is about U.S.$150 per tonne of polluted soils (amount of contaminated soils: 200 ⫻ 10 6 t and estimated costs U.S.$30 billion). It is noted that general costs vary between U.S.$50–500 per tonne of soil, but it may reach U.S.$1000 U.S.$ if complex contaminated soil is in question. In recent years, selected plant species that have the ability to accumulate high concentrations of metals and organic compounds in their roots and shoots from polluted soils have been introduced, especially in the Western world, and this is known as phytoremediation technology. This technology is becoming popular for the clean-up of heavy metals by successive cutting of these plants. This green technology is less expensive and friendlier to nature. However, limited information is available regarding its costs. A few cost results are mentioned in

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Sec. 9. This technology claims that sometimes biomass ash product can be used as a raw material for metals recovery, especially Ni and Co. It is not possible to successfully remove all heavy metals by applying phytoremediation technology, especially Pb. Uptake by plants for Pb is limited. The aim should be to find or develop plants that can remove 500–1000 kg of metal ha ⫺1 yr ⫺1. Researchers have developed tailor-made biosolids using composts rich in hydrous Fe and Mn oxides, phosphates, limestone, clay minerals, wood-ash, and zeolite that can change soil-chemical properties so that contaminants are stabilized, immobile, and not available to plants. Highly effective revegetation was reported by applying phytoremediation technology at Palmerton, Pennsylvania, Katowice, Poland; Bunker Hill, Idaho; and Leadville, Colorado. But phytoremediation also has its drawbacks. It is a time-consuming process, and it may take several years to clean up a site. There is also very limited information regarding the behavior of pests to phytoremediation plants and the depth to which plants can sink their roots to clean up contamination. However, sustainable soil quality and soil protection should be one of the main aims for the government in each country in the new millennium.

ACKNOWLEDGMENT My sincere gratitude is expressed to Professor Dr. Alina Kabata-Pendias, Institute of Soil Science and Plant Cultivation, Pulawy, Poland; Professor W. H. Rulkens, Wageningen Agricultural University, The Netherlands; Dr. R. L. Chaney, USDAARS, Environmental Chemistry Laboratory, Beltsville, MD, USA; Dr. Anna Sophia Knox, University of Georgia, Aiken, South Carolina, USA; and Fredericks Scott, U.S. EPA for supplying me with valuable documents for this project. The assistance received in preparation of this manuscript from the staff members of the Department of Limnology and Environmental Protection, University of Helsinki is gratefully acknowledged.

REFERENCES 1. K Killham. Soil Ecology. Cambridge, UK: Cambridge University Press, 1996, p. 229. 2. JO Nriagu. A global assessment of natural sources of atmospheric trace metals. Nature (London) 338:134–139, 1989. 3. JO Nraigu, JM Pacyna. Quantitative assessment of worldwide contamination of air, water and soils with trace metals. Nature (London) 333:134–139, 1988. 4. E Wahlstro¨m, E-L Hallanaro, S Manninen. The Future of the Finnish Environment. Helsinki: Finnish Environment Institute, 1996, pp. 109–165. 5. WH Schroeder, J Munthe. Atmospheric mercury—an overview. Atmos Environ 32(5):809–822, 1998.

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Biodiversity Index

Abies, 371, 374, 375, 376, 377, 378 Acacia mangium, 304, 305 Acaulospora, 48 Acer sp. 41, 42, 369, 370, 371, 372, 373, 374, 375, 377, 378 Acer pseudoplatanus, 330 A. saccharum, 368 Achanthes microcephala, 70 A. minutissima, 70, 75 Achnanthes, 71, 74, 77 Achnanthes linearis, 70 Acinetobacter, 6 Aeollanthus subacaulis var linearis, 198 Aesculus, 369, 372 Agrobacterium, 6 Agrostis, 163, 459 Agrostis canina, 197

A. capillaries, 41, 193, 196, 197 A. cuneifolium, 166 A. gigantea, 193 A. stolonifera, 197 A. tenuis cv parys, 459 Ailanthus, 369 Albatrellus ovinus, 49 Alcaligenes eutrophus, 6, 9, 29, 30, 31, 33, 35 A. faecalis, 14 Allium cepa, 189 Alnus, 338, 369, 373 A. americanum, 220 A. bertlonii, 193, 220, 225, 228, 229, 229, 230, 240 A. heldreichii, 228 A. lesbiacum, 229, 229, 238, 238 A. malacitanum, 229 473

474

A. mangium, 305 A. montanum, 229 A. murale, 225, 229 A. obovatum, 220 A. perenne, 220 A. pintodasilvae, 229 A. rubra, 418 A. serpillifolium, 229 A. troodii, 228 Alyssum, 220, 223, 225, 228, 229, 230, 238, 240, 458 Amanita, 49, 50 Amanita muscaria, 42, 50 Anabaena, 67 Anabaena cylindrica, 63 A. doliolum, 79 A. flos-aquae, 67 A. inaequalis, 62, 66, 67 A. lutea, 62 A. variabilis, 67 Anacystis nidulans, 61, 62, 63, 67, 74 Andropogon gerardii, 40 Ankistrodesmus, 69 Ankistridesmus falcatus, 64, 83 A. falcatus var mirabilis, 72 Anthoxanthum odoratum, 197 Anthriscus cerefolium, 199 Anthyllis vulneraria, 192 Anthyrium filix-femina, 166 Aphanocapsa sp., 62 Arabidopsis, 171, 185, 293, 294, 302, 457 Arabidopsis thaliana, 184, 312, 313, 455 Arenicola christata, 278 A. marina, 278 Armeria maritima, 191 Arthraxon quartinianus, 198 Arthrobacter, 6 Arthrocnemum fruticosum, 97, 99 Asplenium sp., 166

Biodiversity Index

Asplenium adiantum-nigrum, 166 A. cuneifolium, 166 A. hybrida, 166 A. limneticus, 73 A. presolanense, 166 A. ruta-muraria, 166 A. septiontrionale, 166 A. trichomanes, 166 A. tripolium, 106 A. viride, 166 Asterococcus, 69 Atriplex portulacoides, 278 Bacillus sp., 14, 241 B. subtilis, 19 Barbula recurvirostra, 163 Batrachosperum vagum, 72 Becium homblei, 192, 198 Beilsch miedia, 370 Beta vulgaris, 188 Betula, 41, 42, 330, 331, 373, 374, 378, 380 Betula alleghaniensis, 328 B. papyrifera, 330 B. pendula, 336 B. pendula, 42 B. populifolia 385 B. pubescens, 42 B. tauschii, 361 Biscutella laevigate, 192 Brachymenium acuminatum, 162 B. philonotula, 162 Brassica, 171, 221, 223 B. campestris, 221 B. chinensis, 221, 242 B. juncea, 221, 222, 230–232, 240– 243, 245, 333, 454, 459 B. napus, 244 B. narinosa, 242 B. nigra, 221 B. pekinensis, 221

Biodiversity Index

B. rapa, 221 Bryoria fuscescens, 126, 134 Bryum, 162 B. arachnoideum, 162 Bryum argenteum, 163 B. rubens, 163 Bulbostylis pesudoperennis, 198 Bumilleriopsis filiformis, 81 Burkholderia, 6 Cajanus cajan, 183, 205 Calluna vulgaris, 42 Campylopus bequartii, 162 Cantharellus tubaeformis, 49 Carex, 259 Carpinus, 370 Carya, 369, 371, 372, 373, 377 Castanea, 372, 373, 376 Cedrus, 369, 370, 376 Cephalozia bicuspidata,163 C. hampeana, 162 C. integerrima, 162, 167 C. masalongi, 162, 168 C. nicholsonii, 162, 167 C. rubella, 162 C. stellulifera, 162 Cephaloziella, 162 Cephaloziella calyculata, 167 Ceratophyllum demersum, 267, 276 Ceratopteris cornuta, 165 Cercidiphyllum, 369 Chamaecyparis, 369, 373, 376 Chamaecyparis thyoides, 383 Chamaesiphon minutus, 68 Chamaesiphon subglobosus, 68 Chantarellus ciliarius, 49 Chara, 79 Chara corralina, 63, 83 Cheilanthes inequalis var. lanopetiolata, 165 Cheilanthus hirta, 165, 166 Chlamydocapsa, 69, 71, 74

475

Chlamydocapsa bacillus, 69 Chlamydocapsa cf. petrify, 69 Chlamydomonas acidophila, 70 C. ampla, 79 C. bacilus, 69 C. botryopara, 72 C. debaryana var. micropapilli, 72 C. hebes, 72 C. heterogama, 73 C. kessleri, 62 C. reinhardtii, 65, 79, 80, 81 C. vulgaris, 71, 79 Chlorella, 63, 65, 69, 74, 79, 81, 83 Chlorella fusca, 66, 67, 81 Chlorella fusca var fusca, 72 Chlorella fusca var vacuolata, 69, 72 C. pyrenoidosa, 61, 62, 63, 66 C. vulgaris, 61, 63, 64, 79, 83 C. vulgaris var vulgaris, 72 Chlorococcum, 417 Chroococcus paris, 66 Cinnamomum, 369 Citrobacter freundii, 17 Cladonia, 122 Cladophora, 60, 61, 64, 79 Cladophora glomerata, 60, 66, 67, 69 Comamonas, 6 Closterium striolatum, 72 Cortinarius, 49, 50 Cosmarium, 69, 77 Cryptomeria, 369, 370, 371, 372, 373, 374, 376, 377 Cryptomeria japonica, 357, 364, 379 Cucumis sativus, 186 Cyanidium caldarium, 62 Cymbella, 70 Cylindrocapsa, 69 D. muscorum, 143, 145 Dacrycarpus, 375

476

Datura innoxia, 186, 187 Daucus carota, 312 Dermocybe, 49 Deschampsia caespitosa, 193, 196, 197 D. flexuosa, 197 Desulfobulbus, 4 Desulfomicrobium, 14 Desulfovibrio, 4 Dictyococcus, 69 Diploschistes muscorum, 143 Ditrichum, 162 D. acidophila, 79 D. cornubicum, 162 D. cornubiensis, 167 D. plumbicola, 162, 168 Dunaliella bioculata, 83 D. salina, 87 Escherichia coli, 12, 13, 18, 20 E. hirae, 19 Eichhornia, 259 E. crassipes, 271 E. eximia, 342 Eiseniabicylis, 62 Elsholtizia haichowensis, 198 Enterobacter cloacae, 14 Entorphospora, 48 Eucalyptus, 332, 338, 342 Eucalyptus camaldulensis, 342 Euglena, 79 Euglena gracilis, 66, 73, 79, 81 E. mutabilis, 70 Eunotia, 70 Eunotia exigua, 70, 74 Evernia prunastri, 118 Fagus, 369–372, 374–378 Fagus japonica, 361 Fagus sylvatica, 326, 361, 380 Festuca ovina, 197 F. rubra cv Merlin, 459

Biodiversity Index

F. rubra, 196, 197, 328 Flavoparmelia baltimorensis, 138 Fragaria crotonensis, 81 Fragilaria, 77 Fraxinus, 369, 370, 372, 373, 375, 376, 383 Funaria hygrometrica, 163 Geminella interrupta, 77 Gigaspora, 48 Gingko, 369 Gleditsia, 370 G. turfosa, 70 Gleococcus, 69 Gleocystis gigas, 79 Glomus, 44, 45, 49, 40, 48 Glomus mosseae, 44 Glyceria fluitans, 274, 275 Glycine max, 185, 187, 191, 213, 317 Gomphenema, 70 Gomphidius sp., 49 Gomphospheria sp., 77 Grimmia atrata, 162 Gymnocolea acutiloba, 163 H. macrophylla, 305 H. rivulare, 74 Halimione portulacoides, 97, 99, 107–112 Haumaniastrum katagense, 197, 198 H. robertii, 197, 198 Hebeloma, 50, 331 Helianthus annuus, 192 Holcus lanatus, 196, 197 Hordeum vulgare, 186, 330 Hormidium sp., 71 Hormidium rivulare, 69, 72, 74 H. pseudostichococcus, 72 H. scopulinum, 72 Hovenia, 376 Hybanthus floribundus, 225

Biodiversity Index

Hydnum, 49 H. macrophylla, 303 H. macrophylla, 304, 305 Hydrilla verticillata, 276 Hydrodictyon reticulatum, 61 Hymenoscyphus, 42, 39 Hymenoscyphus ericae, 42, 50 Hypnomonas, 69 H. chlorococcoides, 73 H. chlorococcoides var incrassata, 72 Hypogyminiaphysodes, 133 Idesia, 369 Indigofera setiflora, 192 Ipomea aquatica, 279 Juniperus, 373, 374, 375 Kalopunax, 372, 373 Klebsormidium, 62, 69 K. klebsii, 78 K. rivulare, 70 Laccaria laccata, 49 Lactarius, 49, 50 Lactuca, 171 Larix, 369, 370, 376, 377, 378 Leccinum sp., 49 Lemna, 259 Lemna minor, 199, 277 Leptospirillum, 24 Lindera, 376 Liquidambar, 376 Liriodendron, 376 Liriodendron tulipifera, 328 Liriodendron, 371, 373, 374, 375, 377, 378 Lithocarpus, 369 Lobaria pulmonaria, 129, 130, 131, 132 Lolium multiflorum, 199

477

Lupinus luteus, 186, 187 Lycopersicon peruvianum, 193 L. esculentum, 179, 188, 190 M. cajuputi, 305 M. malabathricum, 304, 305 M. pachyderma, 69 Medicago sativa, 40, 199 Melaleuca cajuputi, 304, 305 Metasequoia, 376 Microcystis aeruginosa, 66 Microspora floccosa, 71 M. pachyderma, 69, 70, 71, 72 M. stagnorum, 69 M. stagnosum, 71, 72 M. tumidula, 69, 71, 73 M. willeana, 69, 71 Microthamnion, 71 Microthamnion kutzingianum, 71, 72 Microthamnion strictissimum, 70, 72 Microthlaspi, 221, 223 Mielichhoferia, 161, 162 M. elongata, 161, 166, 167 M. macrocarpa, 162 M. mieilichhoferi, 162 M. nitida, 161 Mimulus guttatus, 186, 187 Minuartia hirsuta, 177 M. verna, 191, 198 Mohria lepigera, 165 Monoraphidium dybowski, 67 Mougeotia, 69, 71, 77 Mougeotia gracilima, 72 M. parvula, 72 Myriophyllum spicatum, 267 Najas, 259 Nardia scalaris, 163 Navicula, 70, 71 N. pelliculosa, 81

478

Nicotiana plumbaginifolia, 189, 190, 213, 312 N. tabacum 186, 307 N. tabacum L cv Samsun, 307 Nitschia 70, 71 Nitschia closterium, 94 Nitschia palea, 70 Noccaea, 221, 223 Nostoc, 61 Nostoc calcicola, 63 Nothalaena marantae, 166 Ocimum basilicum, 181 Oedogonium, 62, 72, 77 Oidiodendron, 39 Oidiodendron maius, 42 Oligotrichum hercynicum, 163 Oocystis, 69, 71 Oocystis elliptica, 74 O. lacustris, 69 O. nephrocytioides, 69, 74, 77, 79 O. parva, 73 Ophioglossum lancifolium, 166 Oryza sativa, 185, 189, 192, 304, 305 Oscillatoria, 68, 77 P. syringaevartomato, 16 Parmelia caperata, 122, 129- 132 Paulownia, 369 Paxillus involutus, 49 Pellea calomelanos, 165, 166 Peltigera canina, 129, 130, 132, 138 Pezizella ericae, 42 Phacotus lenticularis, 77 Phaseolus aureus, 189, 190 P. banksiana, 179 P. coccineus, 179, 185, 188, 189, 190 P. vulgaris, 177, 179, 180, 183, 184, 186, 188, 189, 190, 199, 328

Biodiversity Index

Phellodendron, 369, 376 Pholianutans, 163 Phormidium, 68, 74 P. foveolarum, 68 P. luridum, 68 P. uncinatum, 68 Photobacterium phosphoreum, 26 Phragmites australis, 274 P. communis, 192 Picea, 370, 371, 372, 373, 374, 375, 376, 377, 378 Picea abies, 49, 325, 328, 381, 382 Pinnularia, 70 Pinnularia acoricola, 70 Pinus, 41, 369, 370, 371, 373, 372, 373, 374, 375, 376, 377, 378 Pinus densiflora, 361 P. nigra, 42 P. ponderosa, 368 P. strobus, 328 P. sylvestris, 42, 46, 380, 386 P. taeda, 328 P. thumbergii, 361 Pisolithus tinctorius, 44, 50, 342 Pisum sativum, 188, 189, 189, 190, 310 Platanus, 369 Plectonema, 68 P. boryanum, 67 Plerococcus, 69 P. rufescens, 72 P. vulgaris, 72 Pohlia andalusica, 167 Pohlia nutans, 163 Polycarpea spirostylis, 198 Polygonum sachalinense, 304, 305 P. deltoides x nigra, 335, 337 P. maximowiczii, 336 P. nigra x maximowiczi, 337, 338 P. nigra, 335 P. trichocarpa x deltoides, 335, 337, 341

Biodiversity Index

Populus, 328, 332, 369, 370, 375, 376 P. tremuloides, 385 Potomogeton, 259 P. pectinatus, 267, 277 P. perfoliatus, 267 Pottia, 162 Prunus serotina, 42 P. virginiana, 328 Prunus, 369, 370, 371, 372, 373, 376, 378 Pseudoanabaena catenata, 77 Pseudococcomyxa, 69 Pseudococcomyxa adhaerens, 71, 72 Pseudomonas aeruginosa, 9, 14, 18, 19, 303 Pseudomonas, 6, 12, 13, 18, 19, 241 Pseudomonas putida, 24, 80 Pseudotsuga, 369, 373, 374, 377 Pteris vittata, 165 Quercus, 139, 369–378, 383 Q. nigra, 368 Q. robur, 368 Q. rubra, 300 Q. velutina, 368 Ralstonia, 25 Ralstonia eutropha, 6, 9–24 R. metallidurans, 6, 9 R. eutropha, 5 Ramalina duriaei, 145 Ramalina farinaceae, 129–132 R. fastigata, 126, 127, 128, 132– 137, 139, 140, 148 Ranunculus, 259 Raparia, 223 Raphanus, 171 Raphanus sativus, 179, 199 Rattus novervegicus, 16 Rauvolfia serpentina, 185, 186, 193

479

Rhizobium leguminosarum bv trifolii, 49 Rhizopogon sp. 50 Robinia, 369, 371 Robinia pseudoacacia, 42 Ruppia, 259 Russula, 49, 50 Sachharomyces cerevisiae, 11, 19, 31 S. bacilaris, 63 Salix, 41, 327, 328, 331, 332, 333, 342 S. arenaria, 335 S. burjatica cv aquatica, 327 S. caprea, 331 S. dasyclados, 339 S. phylicifolia, 42 S. triandra, 339 S. viminalis, 335, 336, 339 Sassafras, 375, 376 Scapania undulata, 165 Scenedesmus, 69, 77, 79, 185 Scenedesmus acutiformis, 69, 79, 81 S. acutus, 66 S. armatus, 73 S. obliquus, 61, 64 S. quadricauda, 81 S. serratus, 73 S. subspicatus, 60, 61, 81 Schizothrix, 68 Scirpus lacustris, 275 Scleroderma flavidium, 330 Scleroderma sp., 50 Scopelophila, 161 Scopelophila cataractae, 162 S. ligulata, 161 Scopelphila cataractae, 161, 167 Scutellospora, 48 Sebertia acuminata, 197, 324 Secale cereale, 184 Selenastrum, 69

480

Selenastrum capricornutum, 63, 65, 74 Senecio cornatus, 225 Serbertia acuminata, 324 Silene cucubalus, 177, 189, 191, 198 S. dioica, 330 S. italica, 185 S. vulgaris, 193, 197 Snechococcus, 81 Solanum lycopersicum, 199 Solidago canadensis, 199 S. aureus, 11, 13, 17 S. typhimurium, 9, 11, 18 Sorbus, 376 Spartina alterniflora, 192, 278 S. maritima, 97, 99, 107–112 S. townsendii, 278 Spinacea oleracea, 179 Spirogyra, 62, 69, 71, 79 S. nitida, 72 Spirulina platensis, 62 Spondylosum pygmaeum, 72 Staurastrum punctulatum, 72 Stewartia, 376 Stigeoclonium, 69, 80 Stigeoclonium aestivale, 69 Stichococcus, 63, 67, 69 Stichococcus bacillaris, 62, 63, 65, 66, 81 Stigeoclonium sp., 71, 80, 81 Stigeoclonium tenue, 69, 71, 72, 74, 81 Streptanthus polygaloides, 458 Styphimurium, 11 Suillus, 50 S. bovinus, 46 S. luteus, 42, 43, 44 Surirella angustata, 70 Synechococcus, 16, 22 Synechocystis aquatilis, 62, 63, 64, 67, 68

Biodiversity Index

Synechocystis sp. 11, 12, 13, 19, 21 Synedra, 70 Synedra filiformis, 70 Syringa, 370 Tabellaria fenestrata, 77 Taxodium, 369, 377 Thalassiosira pseudonana, 81 Thelephora terrestris, 46 Thiobacillus, 3, 4, 6, 24 T. ferrooxidans, 3, 7, 24 Thlaspi, 171, 221, 223, 243, 323 T. alpestre, 223, 233, 234, 235 T. arvense, 236, 238, 239, 245 T. caerulescens, 222, 223, 228, 231, 233, 235, 236, 237, 239, 243, 244, 245, 252, 256, 257, 454, 458 T. calaminare, 233 T. cepaefolium, 233, 237 T. goesingense, 223, 233, 234, 235, 237, 238 T. montanum, 225, 238, 458 T. montanium var calofornicum, 235 T. montanum var fendleri, 238 T. montanum var montanum, 238 T. montanum var siskiyouense, 225 T. montanum californicum subsp. siskiyouense, 235 T. ochroleucum, 181, 239 T. praecox, 223, 223 T. rotundifolium, 223, 237 T. rotundifolium subsp cepaefolium, 223, 233, 234, 235, 245 T. sylvestre subs. calaminare, 239 Thlaspietum cepaefolii, 235 Thuja, 370, 375 Thujopsis, 372, 373 Tilia, 369, 370, 371 Trebouxia, 131 Tribonema, 61, 62

Biodiversity Index

Triticum aestivum, 183, 186, 187, 192, 308 T. durum, 179 T. vulgare, 189, 191 Trochiscia, 69 Tsuga, 370, 374, 375, 377 Typha, 259, 275 Typha latifolia, 192, 274 Uca pugnax, 278 Ulmus, 369, 370, 371, 377, 383 Ulothrix, 61, 69 Ulothrix moniliformis, 69 Ulothrix gigas, 62 U. variabilis, 72 Umbilicaria muhlenbergii, 126, 130 Usnea sp., 126, 127, 128, 136, 137, 139, 140, 144, 148

481

Vaccinium macrocarpon, 304, 305 V. myrtillus, 42 Vallisneria americana, 276 Vaucheria, 62, 63 V. compacta, 81, 82 V. debaryana, 81, 82 Vibrio, 27 Vigna ungiculata, 179, 192, 418 Viola calaminaria, 40, 198 Volerinella locusta, 199 Xanthoriaparietina, 125, 130, 135, 143, 144, 145 Zannichellia paulstris, 267 Zea mays, 179, 191, 199 Zelkova, 376 Zygogonium ericetorium, 71, 72

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Subject Index

ABC (⫽ ATP-binding cassettes) transporters, 10 Accumulation of aluminum, 295 metals by algae, 60 Acid deposition, 379 soils, 289 Adaptations of woody plants to heavy metals, 330 Adverse impacts of metals on soil biodiversity, 415 Alfisols, 289 Algal communities, 68 Aluminum: accumulation, 305, 309 cytotoxicity, 309

[Aluminum] enhanced peroxidation of lipids, 310 excluders, 305 inducible genes, 313 stress, 289 tolerance, 292, 293, 301 tolerant cell lines, 311, 312 toxicity, 289, 295 toxicity and cell wall, 299 toxicity in acid soils, 291 AM fungi, 50 Andisols, 289 Anion-induced cation sorption, 405 Anthropogenic activities, 3 Antiporter systems, 9 Arbuscular mycorrhiza, 38, 40 483

484

Archaebacteria, 6 Atmospheric changes in the world, 365 Bacteria, 1 Bacteria immobilize composite membrane reactor (BICMER), 25 Bacterial heavy metal resistance mechanisms, 7 bark pockets, 353, 354 Behaviour of heavy metals, 433 Beneficial uses of trees in soil remediation, 332 Biological interactions and transitions, 413 oxygen demand (BOD), 335 BIOMET sensors ⫽ Biological metal sensors, 26 Biomonitoring, 385 Biosensors, 28 Biosolid, 460 Blue-green algae (Cyanophyceae), 68 BMSR ⫽ Bio Metal Sludge Reactor, 25 Beneficial effects of aluminum, 303 Calamine soils, 40 Callose formation, 299 production, 309 Carbon mineralization, 417 Cation exchange capacity (CEC), 291, 326, 407 Cation binding capacity of wood, 381 Cation diffusion facilitation, 11 Cation diffusion facilitator (CDF), 11 Cation/Proton antiporter, 9, 18

Subject Index

Cellular growth conditions, 307 Chr ⫽ chromate resistant determinant, 9 Clean-up technologies (Ex-situ extraction), 447 Clean-up technologies (in-situ extraction), 450 cnr ⫽ cobalt, nickel resistant determinants, 9 Coevolution, 45 Complexation, 15 Concentration factor, 60 cop ⫽ copper resistant determinant, 9 czc ⫽ cadmium, zinc and cobalt resistant determinants, 5 Decontamination of soil and water, 24 Dendroanalysis, 367, 368 Dendroanalytical biomonitoring, 386 Development of metal-resistant/ tolerant communities, 74 Diatoms, 70 Diversity, 416 Diversity of algal communities, 68 mycorrhizal fungi, 47 ECM, 45 Ectomycorrhizas, 37, 38, 41 Effect of heavy metals on woody plants, 326 Effect of the sapwood/heartwood boundary on trace element distribution, 382 Effect on microbial processes, 417 Effects of metals on algal growth and metabolism, 64 Efflux, 11

Subject Index

Electrodialytic remediation method, 449 Endomycorrhizas, 38 Entisols, 289 Environmental monitoring, 26 Enzymatic reduction, 13 Ericoid mycorrhizal fungi (ERM), 39 Ericoid mycorrhizas, 42 Fast-growing trees in phytoremediation, 340 Fate of metal ions in the soil environment, 403 Forest ecosystems, 323 Freshwater algae, 59 Functions of mycorhizal fungi, 44 Gene-based biosensors, 26 Genes responding to aluminum, 301 Gram-negative and gram-positive bacteria, 6 Green algae, 69 Halophytes, 112 Hard and soft acid base theory (HSAB), 407 Heartwood polyphenols, 383 extractives, 383 Heavy metal: biosensors, 26 homeostasis, 1 interactions in soils, 401 resistance, 6, 17 tolerance in mycorrhizal fungi, 42 toxicity to microbial communities, 416 uptake, 61

485

[Heavy metal] resistant bacteria, 6 soil compartments, 440 Historical monitoring, 364 Histosols, 289 Hyperacccumulator, 2, 457 Inceptisols, 289 Inhibition of cell division, 297 Inhibition of root elongation, 294 Kaolinite, 404 Land-farming technique, 339 Lead resistant determinant (pbr), 9 Lewis acid, 407, 454 Low molecular weight organic acids, 411 Mechanisms of aluminum tolerance, 312 bark pocket formation, 354 Mercury resistant determinant (mer), 9 Metal: accumulation in roots, 107 avoidance, 79 binding compounds, 80 compartmentalization, 83 contaminated niches, 330 cycling, 112 effects on algal communities, 71 immobilization using biomolecules, 26 leaching, 24 polluted soils, 40 resistance/tolerance, 78 rich anthropogenic biotopes, 3 rich biotopes, 2 rich concretions, 105 sensing, 26

486

[Metal] tolerance, 80 transformations, 83 tolerant plants, 459 interactions vs solid phase composition, 409 Metal fusion proteins (MFPs), 10 Microbial: communities 329 transformation of OM, 414 Mucilage, 302 Mycorrhizal: endophytes, 44 fungi, 37, 39, 40, 41, 43, 45 symbiosis, 37, 40 Natural metal-rich biotopes, 2 Natural sinks for metals, 96 Near-isogenic lines, 294 Nitrogen transformation, 418 Organic acids, 303 Outer membrane: factors, 10 protein, 10 Oxisols, 289 polyaromatic hydrocarbons (PAHs), 25 Pb 210 dating, 363 Periphyton communities, 77 Peroxidation of lipids, 309 pH regulation, 302 Physiology of aluminum toxicity, 299 Phytoextraction, 452 Phytoremediation, 452 Phytostabilization, 452, 459 Phytovolatilization, 455 Plant responses to aluminum stress, 306

Subject Index

Plant uptake of heavy metals, 442 Pollution time capsules, 353, 362 Pollution-induced community tolerance (PICT), 74 Precipitation, 15 Protein based biosensors, 28 Regulation of: ATPases, 21 enzymatic reduction, 23 metal sequestration process, 23 Remediation: polluted soils, 444 principles, 445 Removal of heavy metals, 336 Rhizofiltration, 452 RND ⫽ resistance/nodulation/division, 10 Salt marshes, 95 Sediment chemistry, 101 Selection pressure, 39 Sequestration, 15 Sewage sludge, 51 Short rotation forestry, 333 Sites of metallurgical industries, 4 Soil: bioremediation, 25 enzymes, 419 microbial biodiversity, 401 quality assessment, 50 Sources of metals in soil environment, 402 Spodosols, 289 Sulfate-reducing bacteria, 4 Suspension-cultured cells, 307 Tagus estuary, 97 Thallium-resistant determinant (tll), 9

Subject Index

Thermal treatment of Hg-contaminated soils, 448 Transport and deposition of trace metals in the xylem, 381 Tree: bark, 353 crops, 321 rings 367

487

Ultisols, 289 Vegetation filter, 335 Water use efficiency, 336 Willows and poplar trees, 339 Zinc desert, 41