Fluxes and Effects in Terrestrial and Aquatic ...

Local Organising Committee (LOC) Conference chair: George Gobran Conference Co-Chair: Olle Selinus Conference Secretary: Roger Finlay Göran Ågren, Lars Bergström, Gunnar Jacks, Maj-Britt Johansson, Mats Olsson, Ingrid Öborn Members of the Local Scientific Committee (LSC) All the LOC members and the following experts Bert Allard (ÖU), Hans Borg (SU). Agneta Oskarsson (SLU), Maria Greger (SU), Rolf Hallberg (SU), Kjell Johansson (SLU), Björn Öhlander (LUTH), Klas Rosén (SLU) LOC Assistants Gunilla Jansson, copy editor Fredrik Cederlöf, database Sponsors: KSLA – The Royal Swedish Academy of Agriculture and Forestry, P.O. Box 6806, SE-113 86 Stockholm, Sweden, Ph:+46 8 54 54 77 00, Fax:+46 8 54 54 77 10 Vetenskapsrådet – the Swedish Research Council, SE-103 78 Stockholm, Sweden, ph: +46 8 546 44 000, fax: +46 8 546 44 180 Formas , Swedish Research Council for Environment, Agricultural Sciences and Spatial Planning, Box 1206, 111 82 Stockholm, Sweden, ph: +46 8 775 4000, fax: +46 8 775 4010 Swedish Environmental Agency (Naturvårdsverket), SE-106 48 Stockholm, Sweden, ph: +46 8 698 10 00, Fax +46 8 20 29 25 The Municipality of Uppsala (Uppsala kommun) ,SE-753 75 Uppsala, Sweden, ph: +46 18 727 00 00, fax: +46 18 727 00 01 Nickel Producers Environmental Research Association (NiPERA), 2605 Meridian Parkway, Suite 200, Durham, NC 27713, USA, ph: +1 919 544 7722, fax: +1 919 544 7724 International Council on Mining and Metals (ICMM), 19 Stratford Place, London W1C 1BQ, UK, ph: +44 207 290 4934, fax: +44 207 290 4935 Rio Tinto Health Safety Environment, 5295 South 300 West, Suite 300, Murray, UT 84107, USA, ph: +1 801 743 4628, +1 801 743 4670 International Copper Association, 260 Madison Avenue, 16th floor, New York, NY 10016, USA, ph: +1 212 251 7257, fax: +1 212 251 7245 International Lead Zinc Research Organization (ILZRO), 2525 Meridian Parkway, Suite 100, Durham, NC 27713, USA, ph: +1 919 361 4647, fax: +1 919 361 1957

List of ISTEB Committees ISTEB Corporation Chairman Domy C. Adriano (USA) – chairman Executive Board President- Alex Iskandar (USA) Vice President-Walter Wenzel (Austria) Treasurer-H.Magdi Selim (USA) Secretary-Daniel van der Lelie (USA) Chairperson of the Conference-George Gobran (Sweden) Chairperson of the incoming Conference -Mike McLaughlin (Australia) Chairperson of the past Conference-Les Evans (Canada) Members-at-large: Beverley Hale(Canada), Alina Kabata-Pendias (Poland), Steve McGrath (UK), Bal Ram Singh(Norway) International Committee Nick Lepp (UK) – chair, Catherine Keller (Switzerland) – co-chair for Europe, Herb Allen (USA) – co-chair for North America, Ronaldo Berton (Brazil) – co-chair for Central/South America, Rebecca Hamon (Australia) – co-chair for Australiasia, Ali Boularbah (Morocco) – co-chair for Africa, Bert Allard (Sweden), Alan J. M. Baker (Australia), Amos Banin (Israel), Juan Barcelo (Spain), Nick Basta (USA), Terry Beveridge (Canada), Kim Bolton (Canada), Sally Brown (USA), Ettore Capri (Italy), Hyo-Taek Chon (S. Korea), Thomas Christensen (Denmark), John Cooke (South Africa), Scott Fendorf (USA), Martin Gerzabek (Austria), Carlo Gessa (Italy), Rosanna Ginocchio (Chile), Sabine Goldberg (USA), Carmen Gonzalez Chavez (Mexico), Maria Greger (Sweden), Helja-Sisko Helmisaari (Finland), Philip Hinsinger (France), Shinjiro Kanazawa (Japan), Martin Kaupenjohann (Germany), Anna Karczewska (Poland), Bill Kingery (USA), Kyoung-Woong Kim (S. Korea), Ferenc Laszlo (Hungary), Dar Yuan Lee (Taiwan), Peter Leinweber (Germany), Franz Lobnik (Slovenia), Richard Loeppert (USA), Yongming Luo (China), Lena Ma (USA, Luis Madrid (Spain), Ron McLaren (New Zealand), Wenderly J. Melo (Brazil), Tatyana Minkina (Russia), Jean Louis Morel (France), Ravi Naidu (Australia), Ram Phal Narwal (India), Tamas Nemeth (Hungary), Niels Erik Nielsen (Denmark), Alexander Ponizovsky (Russia), Alexandra Ribeiro (Portugal), David Rimmer (United Kingdom), Brit Salbu (Norway), Milan Sanka (Czech Republic), Henk Schat (Netherlands), Rainer Schulin (Switzerland), A. Paul Schwab (USA), Olle Selinus (Sweden), Nicola Senesi (Italy), Laura Sigg (Switzerland), Eric Smolders (Belgium), Don Sparks (USA), Siobhan Staunton (France), Eiliv Steinnes (Norway), Pavel Tlustos (Czech Rep.), Christos Tsadilas (Greece), Katarzyna Turnau (Poland), Hildegarde Vandenhove (Belgium), Sjoerd E. A. T. M. vander Zee (Netherlands), Jaco Vangronsveld (Belgium), Jean-Frank Wagner (Germany, Wolfgang Wilcke (Germany), Ming H. Wong (China), Satoshi Yoshida (Japan), Scott Young (United Kingdom) Sponsorship/Finance Committee Andrew Green (USA) – Chairperson, Domy Adriano (USA) – member, George Gobran (Sweden) – member, Beverley Hale (Canada) - member, Mike McLaughlin (Canada) member Auditing Committee Harvey Doner (USA) – chairperson, Magdi Selim (USA) – member, Michel Mench (France) – member, Z.S. Chen (Taiwan) – member, Youg-Ming Luo (China) – member

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Messages

Proc. 7th Intern. Conf. on the Biogeochem. of Trace Elements; Uppsala ’03

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Message from the ISTEB Chairman and President It is a great pleasure to welcome everyone to the 7th ICOBTE hosted by the Swedish University of Agricultural Sciences (SLU) in Uppsala. This is again an excellent forum for us and we sincerely thank all the members of the Local Organizing Committee under the leadership of the Chairman (George Gobran), Co-Chairman (Olle Selinus) and Secretary (Roger Finlay) for their hospitality and hard work. On behalf of ISTEB’s Executive Board, the International Committee (Nick Lepp- Chair), Sponsorship/Finance Committee (Andrew Green-Chair), and Auditing Committee (Harvey DonerChair), we invite everyone to again discuss and exchange new and emerging paradigms on the biogeochemistry, bioavailability, and/or risk assessment and management of trace elements in the environment. Indeed, the ICOBTE has emerged as the premier venue for highlighting cutting-edge accomplishments in trace element research that may enhance their beneficial aspects as well as mitigating their detrimental effects in the environment and our society in general. This only became possible because of the commitment and dedication of so many of you from all corners of the world. Together we have sustained and will attempt to elevate our level of enthusiasm for this most challenging endeavor. We have accomplished a great deal of success in forging together a better union with underrepresented groups of the academic/science community, mainly with more women (for example, on the various committees) and more student registration. We have helped focus certain aspects in trace element research, bringing into the arena state-of-the-art spectroscopic and other analytical techniques that have enhanced our knowledge of the behavior of certain elements in various environmental media and organisms at the cellular/molecular level. These frontiers in nanochemistry and nanobiology would undoubtedly aid us to better explore the realm of trace element research never seen before in such fields as surface chemistry, chemical speciation, bioavailability, hyperaccummulation, tolerance/avoidance. While these frontiers may represent one extreme, viewing trace elements at the ecosystem and landscape level could represent the other extreme. The vision that had been seeded earlier by our esteemed mentors, such as our honorees Professors Al Page and Iain Thornton, may just forever serve as our stimulus in our quest to vanquish the most challenging aspects of trace element issues. Synergistic interactions, which ICOBTE may promote among us, may foster a successful journey. May ICOBTE continue its high level of performance to meet everyone’s expectation. And finally, we wish everyone an enjoyable and rewarding conference at SLU and may your journey home be a safe one and hope to see you in Adelaide, Australia in 2005. Domy Adriano ISTEB Chairman and Founder of ICOBTE/ISTEB

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Alex Iskandar ISTEB President and CEO


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Message from the 7th ICOBTE 2003 Chairman On behalf of the Local Organizing Committee (LOC), I would like to express our gratitude and honor in being able to organize the 7th ICOBTE. I must also express our special thank to the International Society of Trace Elements Biogeochemistry (ISTEB) for selecting the Swedish University of Agricultural Sciences (SLU) to host the 7th ICOBTE at Uppsala. The 7th ICOBTE is organized in 14 Scientific Programs and 10 Special Symposia in seven parallel sessions, as well as five internationally prominent keynote speakers. There are over 530 registered participants, including 130 graduate students, from over 70 countries. Special thanks are due to the International Committee (IC) members who devoted considerable time to reviewing over 900 abstracts. We acknowledge the fabulous and effective work performed by Nick Lepp, the chair of the Reviewing Committee that consisted of at least 14 coordinators from the Local Scientific Committee (LSC). The coordination of the Special Symposia by Steve McGrath and Magdi Selim and the organization of the symposia by 10 organizers are highly appreciated. We also thank those who accepted to act as chairs and co-chairs of the conference sessions. Many individuals from Sweden and other countries, who deserve our special thanks and appreciation, have contributed to the accomplishment of this tremendous international activity for little or no remuneration. Generous national and international sponsors have significantly contributed to the success of our 7th ICOBTE. We are particularly grateful to the ILZRO (Andy Green) for financial support. Other important financial support was obtained from research councils and organizations in Sweden, including SLU, FORMAS, SNV, VR and the municipality of Uppsala. The Royal Swedish Academy of Agriculture & Forestry (KSLA) provided generous financial support for a number of scholarships to enable young scientists from the Baltic and neighboring countries to attend the conference. In total the sponsorship we have received has enabled us to subsidies 130 graduate students and to award scholarships to over 30 participants from developing countries. I take this opportunity to thank our Scholarship Committee led by Roger Finlay and Ingrid Öborn for their hard work in overseeing this process. We are thankful to our secretariat staff Tia Eriksson, Birgitta Höglund and Maria Carlson from Akademikonferens, who were responsible for the administration, registration, hotel-booking and website. Many thanks are due to Gunilla Jansson for her efforts in finalizing the copy editing of the two volumes of the conference proceedings, and to Fredrik Cederlöf, who devoted much time to the database of the abstracts and responding to a flood of e-mail messages on a daily basis. Special thanks to Göran Ågren for providing invaluable help in planning and editing of the Conference Program. I personally would like to thank all my colleagues in the LOC whose tireless efforts over the past two years have contributed to the success of this conference. All member of the LOC are expressing their indebtedness to the rector of SLU, Ann-Christin Bylund for her constant support and interest in making the 7th ICOBTE successful. George Gobran SLU, Uppsala, June 2003


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Keynotes


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Transfer of Trace Elements through the Food Chain Mike McLaughlin

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CSIRO Land and Water, PMB 2, Glen Osmond, Adelaide, 5064, AUSTRALIA (E-mail: Mike.McLaughlin@csiro.au) 2 Soil and Land Systems, The University of Adelaide, Adelaide, 5064, AUSTRALIA At low concentrations some trace elements (e.g. copper, chromium, molybdenum, nickel, selenium, zinc, etc.) are essential to healthy functioning and reproduction of microorganisms, plants and animals (including man). However, at high concentrations these same essential elements may cause direct toxicity or reproductive effects. Some trace elements are also non-essential (e.g. arsenic, lead, mercury, etc.) and even low concentrations of these elements in the environment can cause toxicity to both plants and animals. Adverse effects are not necessarily only manifested in the environment when trace elements have an anthropogenic origin – naturally high concentrations of some elements also cause toxicity (e.g. serpentine soils) and lead to natural adaptation of the biota to these elevated concentrations. Food chain effects are manifested when bioaccumulation and biomagnification of elements occurs across trophic levels, so that organisms at higher trophic levels are threatened by concentrations harmless to lower organisms. In the terrestrial environment, there are several barriers to the bioaccumulation and biomagnification of trace elements. In uncontaminated environments, the bioavailability of many trace elements is often low, due to the presence of highly insoluble or occluded forms in soil. Hence, transfer through the food chain is limited by low concentrations in soil solution. Where soil pollution occurs, the bioavailability of added trace elements may also be very low depending on the pollution source, e.g. in silicate slags or insoluble oxides, thus minimising food chain risks. If trace elements are added to soils in a soluble, and hence highly bioavailable form, reactions with soil surfaces and solutes may also render insoluble precipitates or strongly adsorbed species, also keeping risks from food chain transfer low. Lead (Pb) is a good example of such an element, where total, or added concentrations, are hazardous, but in fact are low food chain risk when fate and transport issues are considered. A further issue to be is the propensity for organisms to actively exclude some elements from uptake, or mitigate transfer to organs consumed by higher trophic levels. Chaney (1980) combined the above considerations to suggest “the soil-plant barrier” concept to rank trace elements for food chain risk assessment, and the concept is still valid. Table 1. Trace elements grouped for food chain risk according to the “soil plant barrier concept” (from Chaney, 1980).

Group 1 – solubility limited Ag Cr Sn Ti Y Zr

Group 2 – translocation limited As Hg Pb

Group 3 – phytotoxic B Cu Mn Mo Ni Zn

Group 4 – higher risk Cd Co Mo Se

Of the elements considered as high risk, Cd is perhaps the element which has traditionally raised the most concern, due to widespread contamination of soil by inadvertent addition of Cd to soil in fertilisers and soil conditioners (manures, mulches and recycled urban wastes). The pathways of Cd from soils to plants and animals, and thence to humans is complex (Fig. 1), but real effects on humans

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have been manifested in the past. Food chain biomagnification of Cd has also been demonstrated to have had adverse effects on wildlife. NON-FOOD INTAKE Smoking Occupational exposure

INPUTS Atmosphere Fertilizer Sewage biosolids Soil amendments Industrial wastes Agricultural wastes

Atmospheric contamination & uptake

Translocation

FOOD INTAKE Wast es

Manu re

Grazing

Root

upta

ke

SOIL SOLUTION POOL

ge an ch ex ct ta on C

Free metal + complexed (Cd 2+ + CdL) Adsorption Desorption

EXCHANGEABLE POOL

No M n-a in ss er im al i isa la tio tor n y

LABILE POOL

Fixation

Mobilisation NON-EXCHANGEABLE POOL (PRECIPITATED, CRYSTALLINE, ADSORBED)

NON-LABILE POOL

MINERAL PHASE

ORGANIC PHASE Leaching

Fig. 1. Transfers of Cd in the soil-plant-animal-human system (from McLaughlin and Singh, 1999).

Cadmium is relatively easily taken up by plants, and in soils there are several reactions and situations, relatively common in agricultural systems, which may increase food chain transfer e.g. acidity, salinity, or zinc-deficiency. While much progress has been made in understanding the agronomic management of this element, there is still uncertainty regarding the safe levels of intake by humans which should be used to set food quality benchmarks for good agricultural practices. This issue needs to be resolved to both ensure human health protection and to avoid artificial barriers to global trade in food commodities. REFERENCES Chaney, R.L. (1980) Health risks associated with toxic metals in municipal sludge. In: Bitton, G., Damro, B.L., Davidson, G.T., Davidson, J.M. (Eds.), Sludge - Health Risks of Land Application. Ann Arbor Sci. Publ., Ann Arbor, MI, USA. pp.59-83. McLaughlin, M.J., Singh, B.R. (1999) Cadmium in soils and plants. A global perspective. pp. 1-9. In: McLaughlin, M.J., Singh, B.R. (Eds.) “Cadmium in Soils and Plants”. Kluwer Academic Publishers, Dordrecht, The Netherlands.


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Trace Elements and Soil Protection in Europe Giovanni Bidoglio European Commission, Joint Research Centre, Institute for Environment and Sustainability, Soil and Waste Unit, 21020 Ispra, ITALY (E-mail: giovanni.bidoglio@jrc.it) Trace elements in soils have been subjects of research for decades. The literature abounds with information on fate, availability to plants and organisms, methodologies for detection of trace elements. However, this knowledge has not often been translated into regulatory actions. The lack of driving force was mainly associated to the fact that specific provisions for the soil compartment were not perceived to be needed as it was the case for mobile media (air, water). Additionally, soil is a complex medium of multiple functions lending itself to different approaches, e.g. even the definition of soil is often matter of discussion among soil scientists, which did not help convince regulators and the public. There is now a common concern in Europe regarding soil protection and pollution prevention. A Thematic Strategy for Soil Protection has been proposed also at EU level. Concerning trace elements, a number of initiatives of European relevance have been identified that cover, among others, mining, land spreading of sewage sludge, deposition of airborne trace metals, monitoring. Taking into account that policy initiatives are often triggered by accidents or perceived hazards, should we then consider that soil distress is now exhibiting its negative effects? Regulations for trace elements in soils have developed in Europe mainly for management, clean up and reuse of locally contaminated areas. In this context, guideline-action-threshold-background, etc. values have been introduced. These values have been also attached preventive objectives and helped integrate soil requirements into other policies. However, a critical problem is the terminology adopted to define these guideline values, which may have different meanings in the different countries. This hinders a systematic comparison of soil prevention issues, including policy approaches at European level. The “reference framework” for regulators is represented by total element concentrations considered as hard values easy to obtain. Large investments have been made to carry out Europeanwide measurements of impacts of trace elements in soils. The FOREGS initiative for a Geochemical Baseline for Europe is in its final stage; the ICP Forest will carry out soon a second large-scale survey for the evaluation of the quality of forest soils in the context of Council Regulation on the Protection of Forests against Atmospheric Pollution; the 2000/2001 survey of atmospheric deposition of trace metals in Europe using mosses has involved 26 countries. In addition, all European countries have their own soil monitoring programmes for trace elements. The need for harmonisation and standardisation to optimise the use of resources is evident. Moreover, the concept of bioavailability should be integrated into regulatory systems. Crucial to this end, is the establishment of a common framework through the introduction of standard methods. Regulatory initiatives for sustainable development in Europe need a long-term pragmatic contribution from soil science and trace element research. The EU Soil Thematic Strategy linked with a possible proposal for a European Integrated Environment and Health Monitoring System can provide the framework for such a systematic approach.

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Phytoremediation – Does It Work? Maria Greger Department of Botany, Stockholm university, SE-106 91 Stockholm, SWEDEN (E-mail: maria.greger@botan.su.se) Phytoremediation is defined as the use of green plants to remove, contain, or render environmental contaminants harmless (Cunningham and Berti, 1993). Various phytoremediation techniques exist for different contaminants and medium. In the case of trace elements, following techniques can be presented (Salt et al., 1995); ● phytoextraction – element accumulating plants are established on contaminated soil and later harvested in order to remove the specific elements from the soil; ● rhizofiltration/ phytofiltration — roots or whole plants of element accumulating plants absorb the element from polluted effluents and are later harvested to diminish the metals in the effluents; ● phytostabilisation — plants tolerant to the element in question are used to reduce the mobility of elements, thus, they are stabilised in the substrate or roots. One reason to use plants for clean up is a relative low cost and maintenance requirements (Cunningham and Berti, 1993). The most investigated technique is phytoextraction. The idea probably came up as a consequence of the findings of plants that hyperaccumulate different elements (Baker et al., 2000). A lot of effort has been paid to investigate the efficiency of various hyperaccumulators to accumulate metals in the shoot as well as the mechanisms of uptake, accumulation and translocation to the shoot. The root soil interaction has also been investigated and the removal of metals seems to be most obvious in the bioavailable fraction. Addition of additives such as EDTA increase the availability of some metals and thus also the phytoextraction efficiency. Since hyperaccumulators in general have low biomass production also high accumulators, (plants with lower accumulating capacity than hypraccumulators) with high biomass production have been discussed in phytoextraction (Greger and Landberg, 1999). The reason is that to evolve an efficient removal of an element not only the uptake and shoot accumulation of the element is important but also the biomass, since it is the removal per area that counts. There is a hope to be able to combine the metal accumulating properties of hyperaccumulators with high biomass production in the future to create a fast phytoextracting plant. Calculation on phytoextractions of highly contaminated sites, such as mine waste, based on the metal accumulation in the plants shows that a decrease of the soil metal content to an uncontaminated level would take several hundreds of years. Therefore, phytoextraction of low-medium contaminated soils seems to be more appropriate. In rhizo and phyto filtration of polluted water roots penetrating the water, emerged or submerged plants can be used. To decrease the concentration of specific elements in water plants with high uptake efficiency is to prefer. Also in this case a high biomass per water volume is necessary for efficient water cleaning. Not only uptake but also increase in sedimentation of metal contaminated particles due to the presence of plants helps to remove metals from the water. Establishing vegetation for stabilisation of soil to prevent erosion is a known method. In addition, in phytostabilisation the elements are bound in the rhizosphere or taken up into the roots. A low translocation to the shoot of the specific elements will reduce the risk of element dispersion by grazing animals or leaf senescence. Phytostabilisation seems to be the best functioning method of all phytoremediation methods. It may be a successful way to prevent the formation of acid mine drainage and stabilize metals in mine tailings (Tordoff et al., 2000; Stoltz and Greger, 2002).

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REFERENCES Baker, A.J.M., McGrath, S.P., Reeves, R.D., Smith, J.A.C. (2000) Metal hyperaccumulator plants: A review of the ecology and physiology of a biological resource for phytoremediation of metalpolluted soils. In: Terry, N., Banuelos, G. (Eds.), Phytoremediation of contaminated soil and water. CRC Press, Boca Raton, Florida. Cunningham, S.D. Berti, W.R. (1993) Remediation of contaminated soils with green plants. An overwiew. Vitro Cell. Dev. Biol., 29, 207-212. Greger, M., Landberg, T. (1999) Use of willow in phytoextraction. Int. J. Phytorem., 1, 115-123. Salt, D.E., Blaylock, M., Kumar, N.P.B.A., Dushenkov, V., Ensley, B.D., Chet, I. Raskin, I. (1995) Phytoremediation: A novel strategy for the removal of toxic metals from the environment using plants. Biotech., 13, 468-474. Stoltz E., Greger, M. (2003) Cottongrass effects on trace elements in submersed mine tailings. J. Environ. Qual., 31, 1477-1483. Tordoff, G.M., Baker, A.J.M., Willis, A.J. (2000) Current approaches to the revegetation and reclamation of metalliferous mine waste. Chemosphere, 41, 219-228.


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Bioavailability – A Concept Driven by Science or Legislation? Nicholas W. Lepp School of Biological and Earth Sciences, Liverpool John Moores University, Byrom Street, Liverpool L3 3AF, UK (E-mail: N.W.Lepp@livjm.ac.uk) INTRODUCTION The problems created by metal contamination of soil and input of industrial wastes is a worldwide problem. The presence of toxic metals raises questions relating to land use, ecological and human health risks and issues of liability. The first two aspects are currently decided on largely scientific grounds whilst the latter is based on legal assessment of their consequences. All are based on the concept of bioavailability. BIOAVAILABILITY – EASY TO DEFINE, IMPOSSIBLE TO MEASURE? Metal bioavailability is the fraction of the total metal burden present in a given matrix, which can be utilized by biota. In most cases, this is equated with solubility, although there are exceptions; some microorganisms can extract metals directly from a solid matrix using extracellular secretions. Root exudates can directly and indirectly solubilise metals. Re-entrained particulates can be inhaled or ingested then solubilisd within a target organism. Here, bioavailability is a direct function of particle dissolution in the digestive system or lungs without the intervention of secondary factors. Direct measurement of bioavailability is possible in these circumstances. Problems arise when measuring bioavailability in complex ecosystems where metal solubility is a function of interacting biological, chemical and physical factors. Transfer of bioavailable metals depends on the assemblage and complexity of the biota. Measurements are modified by restricted transfer at each level, making identification of a target organism a fundamental process in estimating bioavailability. Metal bioavailability in soil is indirectly estimated using chemical extractants that remove metals from operationally defined labile fractions. This approach underpins development of phytoextraction strategies, and yet bears no convincing correlation to fractions available to soil microbes and plants. Phytoremediation relies on understanding the rate that bioavailable fractions are replenished from the soil matrix. A generally applicable method to measure bioavailable soil metals must be developed. HUMANS AS TARGET ORGANISMS What are the most important aspects of bioavailability for the protection of human and ecological health? They can be broken down into direct and indirect considerations, based upon exposure pathways. Direct exposure relates to metals in re-entrained particulates and potable water. Risks associated with these are removed when exposed soil is covered by vegetation or measures taken to restrict metal leaching to groundwater. The former depends on bioavailability as it affects plant establishment and longevity. A permanent, self-sustaining vegetation cover over contaminated soil reduces potential for direct human exposure. The latter may be prevented by engineering techniques or chemical treatments. Indirect exposure follows movement of metals into different trophic levels of the human food web. Attention should be focussed on cultivation practices, dietary patterns, and food sources. Bioavailability of metals in all types of agricultural soils, at a level that does not reduce crop quality, should be of primary concern. Crops that accumulate metals in the absence of phytotoxicity, such as lettuce and spinach, form a limited part of most diets. Vegetarian diets have grains or pulses as major constituents, supported by tubers, seasonal fruits and vegetables. These are sourced on a worldwide basis in developed countries, so potential magnification of plant-borne metals in the human system is limited. In developing countries, where dietary constituents tend to be sourced and consumed locally, exposure potential is greater.

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The use of plants that show significant inefficiency in metal uptake is relevant where agricultural soils have become contaminated. The genome of many food crops is variable, with features retained or enhanced as by-products of breeding programmes. Cadmium uptake in lettuce is one example, as is restricted As uptake by Brassicas METAL BIOAVAILABITY AND DEVELOPMENT OF SUSTAINABLE ECOSYSTEMS There are many examples of successful, natural colonisation of metal-contaminated sites by native flora, in historic mining areas and regions at the forefront of the 19th century industrial revolution. In the latter case, inefficient examples of chemical engineering created massive volumes of toxic wastes that were indiscriminately dumped, creating extensive areas of polluted soils. In the post-industrial dereliction that accompanied the decline of such regions, highly polluted soils have been naturally colonised by vegetation, creating stable, sustainable ecosystems that reduce the risk from soil metal pollution. Studies at a site in the UK, reported elsewhere in this meeting, show that over a period of less than 50 years, a stable, functional ecosystem has naturally established on a site that is classified as ‘toxic waste’. Mechanisms underpinning this remarkable transformation are concerned with restriction of metal acquisition by dominant species in the ecosystem, reducing phytotoxicity and preventing significant food chain transfer. SOME CONCLUSIONS Bioavailability, as currently understood, is a concept that is more meaningful to legislators than to scientists. In many respects, the limitation and prevention of human and ecological health risk relies heavily on an imprecise and poorly understood science. The differences that exist in metal bioavailability to plant species growing in the same polluted soil illustrates the problem of transferring the concept to a strict legal framework. In most cases, bioavailability of soil metals has a negligible impact on quality of human life; vegetation can colonise and develop in metal polluted soils, stabilising and covering exposed surfaces and significantly reducing direct exposure risk. However, agricultural soils pose a greater risk when crops grown on polluted soil form a significant part of the human diet. In reality, dietary inputs are so diverse that metals from this source are a minor risk to health. Identification and cultivation of crop plants that exclude metals is a significant method of further reducing risk.


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Trace Element Chemistry, Contamination and Ecotoxicity Gary M. Pierzynski Department of Agronomy, Kansas State University, Manhattan KS 66506 USA (E-mail: gmp@ksu.edu) Trace element chemistry, contamination and ecotoxicity have been the subjects of considerable research for the past 30 years. The purpose of this paper is to qualitatively examine the justification for continuing this effort and to consider future directions for trace element research. A significant part of the justification for soil trace element research has been potential impacts of trace elements on human health. Nriagu (1988) estimated that as many as 1 billion people could be adversely affected by Pb, 500,000 by Cd, 80,000 by Hg, and >100,000 by As. These estimates were made using health survey data, critical values for medical parameters or exposure media, and a variety of assumptions. More recent estimates do not appear to be available. However, judgments can be made as to the relative change in the number of people impacted today based on new information. The number of persons affected by Pb would likely be lower today by the standards used by Nriagu (1988) because of reduced use of leaded petrol worldwide. However, recent data suggest significant human health effects from Pb at levels of exposure lower than originally believed or used by Nriagu (1988), and the number of people affected by Pb today likely remains quite high (Canfield et al., 2003). Human health effects from As have likely increased in the past 15 years as recent occurrences of As poisoning in India and Bangladesh were not considered by Nriagu (1988). Human health effects alone do not drive trace element research. As human health effects have been addressed in some countries, the potential ecological impacts of trace elements have received greater attention. Ecological impacts are often viewed as only impacting species other than humans. However, ecological impacts address the broader issue of our resource base and ecological impacts can and will ultimately impact humans. Ecological impacts are more difficult to address because of multiple species, interactions between species, significant issues with respect to bioavailability in various media, and the large areas that require attention. A number of topics will dominate trace element research in the near future. A major issue in soil science in general is the use of model systems to study chemical processes as opposed to using “real” samples. Progress will continue to be made on studying chemical processes in real samples. An excellent example of this has been increased use of synchrotron-based x-ray absorption spectroscopy in the soil and environmental sciences over the past 10 years. A number of techniques are employed with this general approach and all are very powerful when applied to model systems. The major limitation is that the weighted average signal is obtained from all chemical forms of an element in a sample. Some investigators have attempted to overcome this by employing varying proportions of mineral phases as standards and mathematical fitting algorithms (e.g., Paktunc et al., 2003). Progress has been made on increasing spatial resolution by reducing the beam size, particularly for x-ray fluorescence (e.g., Punshon et al., 2003). Another analytical advancement that will impact trace element research is the reduced cost, and subsequent increased availability, of ICP-MS. Remediation will remain a researchable topic into the foreseeable future, particularly for ecological risk. There are large areas in need of remediation and the demand for low-cost and effective methods remains high. Soil amendments will continue to play a role with additional emphasis on prescription soil amendments for specific situations. An ancillary issue will be natural attenuation as, despite efforts at ecological restoration, there will not be sufficient resources to address all areas in need of remediation. Concurrent with the ecological risk efforts will be research supporting risk-based guidelines and regulations. Bioavailability will also remain as an important research topic in various permutations. The development of laboratory tests as surrogates for animal feeding studies will continue. Animal feeding studies have become the standard for assessing bioavailability of As and Pb in the United

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Keynote

States despite the fact that such studies take considerable amounts of time and are expensive. Surrogate tests would allow, at the very least, a screening method useful for environmental assessments and for evaluating soil treatments designed to reduce trace element bioavailability. The assessment of the bioavailability of soil constituents to plants and soil invertebrates will be another area of active research. Batch extraction methods cannot assess the potential transport of trace elements through the soil and may not be able to assess the appropriate chemical species. Consequently, correlations between extractable trace elements and response in one soil generally do not hold across different soils where different soils would represent variations in soil physical properties and trace element sources. Diffusive gradients in thin films (DGT) and comparable approaches use ion sinks to assess trace element bioavailability and, when used in situ, can overcome some of the shortcomings of batch methods. The use of DGT and related methods will likely increase and this may allow for the development of a more standard set of critical values for toxicity. Finally, the future of trace element research will involve more multi-disciplinary efforts. Recently we have seen soil scientists conducting animal feeding studies and there will be continued collaboration between with toxicologists. Numerous other opportunities exist including work with the engineering, medical, statistical, and biological fields. All indications are that there are many opportunities for the future of trace element research and there should be strong support for the ICOBTE series for many years to come. REFERENCES Canfield, R.L., Henderson, C.R., Cory-Slechta, D.A., Cox, C., Jusko, T.A., Lanphear, B.P. (2003) Intellectual impairment in children with blood lead concentrations below 10 micrograms per deciliter. New Eng. J. Med., 348, 1517-1526. Nriagu, J.O. (1988) A silent epidemic of environmental metal poisoning. Environ. Pollut., 50, 139-161. Paktunc, D., Foster, A., Laflamme, G . (2003) Speciation and characterization of arsenic in Ketza river mine tailings using x-ray absorption spectroscopy. Environ. Sci. Technol., 37, 2067-2074. Punshon, T., Bertsch, P.M. Lanzirotti, A. McLeod, K. Burger, J. (2003) Geochemical signature of contaminated sediment remobilization revealed by spatially resolved x-ray microanalysis of annual rings of Salix nigra. Env. Sci. Technol., 37, 1766-1774.


15

SP01or - Fluxes and Effects in Terrestrial and Aquatic Ecosystems

16



Fluxes and Effects in Terrestrial and Aquatic Ecosystems Scientific Program 1 Oral Wednesday June 18


17


Process Studies on Mercury Fluxes over Various Soils with a New Laboratory Flux Chamber System (LFCS) E. Bahlmann, R. Ebinghaus Institute for Coastal Research GKSS Research Centre, Max Planck-Straße 1, 21502 Geesthacht, GERMANY (E-mail: enno.bahlmann@gkss.de ralf.ebinghaus@gkss.de) INTRODUCTION Conservative estimates of global natural mercury fluxes into the atmosphere suggest a total of 500 t/a originating from the mercuriferous belt and about 200 t/a degassing from background soils. (Lindqvist et al., 1991) Recent flux measurements at different terrestrial sites suggest that mercury emissions and re-emissions from terrestrial sites have been significantly underestimated. (Ebinghaus et al., 1999) This could result in an inaccurate ratio between natural and anthropogenic sources in global emission inventories. However, all this estimates are based on a fairly raw database, and considering that, it becomes clear that an accurate assessment of air/surface exchange processes will be critical to refine the global mercury cycle. For upscaling mercury fluxes derived from field experiments to regional or global scales a detailed process understanding of the driving forces behind air/surface exchange is imperative. MATERIAL AND METHODS For our studies we used a Laboratory Flux Measurement System (LFMS). The system is designed for the simultaneous flux determination of mercury, CO2 and water vapour under controlled environmental conditions. The System operates with two identic flux chambers, one acting as a reference and the other as a sample chamber. The concentrations of the analytes are measured simultaneously at the outlet of both chambers. Fluxes are calculated on the basis of the outlet concentrations. All measurements are corrected for standard conditions. Total gaseous mercury (TGM) is measured with two Tekran Mercury Vapour Analyzers (Model 2537) with a time resolution of 5 min. Concentrations of CO2 and water vapour are determined continuosly using a Two Channel Infrared Gas Analyser (LiCor, Model Li 6262). Air and soil temperature, light intensity and soil humidity are measured continuously on a routine base. The cylindrical chambers have a diameter of 50 cm corresponding to a surface area of ~2000 cm² and a variable height between 10 and 40 cm, resulting in a chamber volume from 20–80 L. All system parts coming in contact with samples are made of or coated with Teflon-FEP Inlet- and outlet ports are on opposite sides of the chamber in a height of 8cm over the sample surface. A fan is installed in the center of each chamber for continous mixing of incoming air. Ambient Air is pulled through the chambers at predefined flow rates between 1.5–30 l/min, using high capacity mass flow controllers and membrane pumps. Water can be added to the soil samples as water vapour or as a fine aerosol via the inlet port of the chambers. An additional port at the bottom of the chambers is designated for direct additon of water or solutions to the soil samples. The bottom of the chambers are designed as a heat exchanger and coupled with a thermostat (LaudaUKT 600) allowing control of the soil temperature from 35 to 10 °C. To investigate the influence of solar radiation on mercury exchange processes we use a solar simulator. RESULTS AND DISCUSSION For all soils under investigation we found a strong influence of radiation, soil-surface temperature and turbulence on the momentum emission flux of mercury. Temperature differences between the soil surface and the air above showed a moderate but still significant influence on the emission fluxes. First results on the influence of soil moisture on mercury emission fluxes show that these two parameters are positively correlated. Clear evidence for a photoinduced increase in mercury fluxes for all soils could be revealed. Comparison of measured and calculated fluxes (Fig. 1) derived from the

18



temperature dependency of the mercury flux (F-TGM = 8.3626e0.0923x, x=T[°C]) showed good agreement for “dark-phases“ but failed for “light“-phases within this time series This observation is probably due to an additional photochemical formation of elemental mercury at the soil surface. F-TGM [pmol/m²h] 200

160

T [°C] 120

F-TGM measured radiation 320 W/ ² F-TGM calculated Soil Surface Temperature 0.5 cm

96

120

72

80

48

40

24

0 5875

0 6125

7035

7285

7535

7785

8035

8285

9140

time [min]

Fig.1. Comparision of measured and calculated Fluxes for a sandy podzolic soil.

Our results suggest that for a given soil variation in mercury emissions fluxes can be explained by variations in soil surface temperature, turbulence and radiation. A comparison of the fluxes obtained for different soils indicate that the differences in fluxes are well correlated with the mercury concentrations in the soils. REFERENCES Lindqvist, O., Johansson, K., Aastrup, M., Andersson, A., Bringmark, L., Hovsenius, G., Hakanson, L., Iverfeldt, A., Meili, M., Timm, B. (1991) Mercury in the Swedish environment –recent research on causes, consequences and corrective methods. Water, Air, and Soil Pollut., 55, 261. Ebinghaus, R., Tripati, R.M., Wallschläger, D., Lindberg, S.E. (1999) Natural and Anthropogenic Mercury Sources and their Impact on the Air-Surface Exchange of Mercury on Regional and Global Scales In: Ebinghaus, R., Turner, R.R., Lacerda de, L.D., Vasiliev, O. Salomons, W. (Eds.) Mercury Contaminated Sites: risk assessment and remediation, Springer, Berlin, Heidelberg, pp. 3 – 50.


19


Speciation and Biogeochemistry of Hg in Two Shallow Estuaries: Lavaca Bay (Texas) and Venice Lagoon (Italy) 1

2

2

N.S. Bloom , L.M. Moretto , P. Ugo 1

Frontier Geosciences Research and Consulting, 414 Pontius Avenue North, Seattle, 98109, USA (E-mail: nicolasb@frontiergeosciences.com) 2 Deptartment of Physical Chemistry, University of Venice, S Marta 2137, Venice, 30123, ITALY (E-mail: moretto@unive.it ugo@unive.it) INTRODUCTION The Hg cycle of shallow estuarine systems is dominated by direct coupling between the sediments and the food chain. There is a high ratio of sediment surface to water volume, increased resuspension, and proximity to intertidal wetlands. We compare occurrence, speciation, and dynamics of Hg in two estuaries, Lavaca Bay (Texas, USA, 1996-1997) and Venice Lagoon (Italy, 2001-2002). Both are large, shallow embayments, contaminated from1950s -1980s by Hg cell chlor-alkali plants. Both have rapid tidal flushing and strong seasonal variations in water temperature. The Venice Lagoon is located in a densely populated industrial/agricultural region (population 1,500,000), while Lavaca Bay is located in a rural agricultural area (population 41,000). Lavaca bay receives about 10 % of its water budget from the Lavaca River, whereas the Venice Lagoon has small direct inputs of fresh water, following historical redirection of all major rivers. MATERIAL AND METHODS Samples were collected and analyzed using ultra-clean methodologies (Bloom et al., 1999). Water was collected by hand dipping into Teflon bottles from the bow of a moving boat, then filtered and preserved at the laboratory within a few hours. Surface sediments were collected into 20 mL glass vials and immediately chilled (0 ºC) until being frozen. Total Hg was determined on sediments after aqua regia digestion and on waters after BrCl digestion using SnCl2 reduction, purge and trapping on gold, and cold vapour atomic fluorescence spectrometry (CVAFS). Methyl Hg was determined using aqueous ethylation, isothermal GC separation, and CVAFS detection. RESULTS AND DISCUSSION Although both sites were contaminated by direct discharges prior to the 1980s, since the advent of pollution control, Hg levels have diminished to moderate levels in most ecological compartments (Table 1). In Lavaca Bay, stable sedimentation has buried the sharply defined Hg maximum (approximately 10-50 ng/g Hg in 1970) at depths of 10-40 cm (Bloom et al., 1999) while at most sites in the Venice Lagoon, Hg concentrations vary only slightly throughout the upper 50 cm (Donazzalo et al., 1984). This is due to a greater degree of anthropogenic disturbance within the Venice Lagoon. The Venice Lagoon shows similar levels of CH3Hg to Lavaca Bay (Table 2). This is somewhat surprising, given the greater area of wetlands in the Venice Lagoon (30 % versus less than 5 %). A dramatic increase in methylation seen in early spring in Lavaca Bay (just at the onset of warming) was not observed in the Venice Lagoon, where the levels of CH3Hg continue to increase slowly throughout the year. In addition, the Hg concentrations appear to be spread far more widely in the Venice Lagoon, possibly resulting in a higher system-wide production, when multiplied by the three times larger surface area.

20



Table 1. Ranges of Hg in sediments (dry basis) and unfiltered water samples.

Sample type

Lavaca Bay Total Hg 0.5-1.0 5-30 10-30 300-2 000

Water, “background” ng/L Water, “contaminated” ng/L Sediment, “background” ng/g Sediment, “contaminated” ng/g

Methyl Hg 0.01- 0.02 0.25-1.1 0.05-0.2 1-10

Venice Lagoon Total Hg 0.5-1.0 7-80 20-50 400-5 000

Methyl Hg 0.01-0.02 0.10-0.35 0.05-0.2 0.5-5

Data from both sites indicate that sediment CH3Hg concentrations are more strongly correlated with variables such as TOC, habitat type, and seasonality than with the total Hg concentrations. No correlation was found between organism CH3Hg concentrations and the very high levels of historic total Hg contamination that are buried 10-30 cm below in Lavaca Bay. This, combined with the observation of massive Hg methylation after dredging activities, argues for leaving Hg contaminated sediment in place, to be buried by deposition of cleaner sediments. In Venice, is this complicated by continuous sediment mixing and the larger area of contamination. Table 2. Speciation of Hg of suspended matter from the shallow zones (< 0.5 m). Because of high suspended solids (4-90 mg/L), about 96% of the Hg and 72% of the MHg at these sites is found in the suspended matter.

Location

Season

N

Lavaca Bay

Winter Spring Summer Winter Spring Summer

2 17 6 2 7 6

Venice Lagoon

Hg (ng/g (dry basis)) Total Methyl 550 ± 164 3.0 ± 0.5 1,301 ± 150 11.5 ± 3.1 1,116 ± 347 14.9 ± 3.8 967 ± 6 4.7 ± 2.3 1,741 ± 886 6.0 ± 1.4 1,145 ± 472 29.1 ± 14.4

% Methyl 0.63 ± 0.28 1.03 ± 0.21 1.58 ± 0.25 0.49 ± 0.23 0.61 ± 0.14 1.77 ± 0.15

REFERENCES Bloom, N.S., Gill, G.A., Cappellino, S., Dobbs, S., McShea, L., Driscoll, C., Mason, R., Rudd, J. (1999) Speciation and Cycling of Mercury in Lavaca Bay, Texas, Sediments. Environmental Science and Technology, 33, 663-669. Donazzolo, R., Orio, A.A., Pavoni, B., Perin, G. (1984) Heavy Metals in Sediments of the Venice Lagoon. Oceanologica Acta, 7, 25-32.


21


Critical Levels of Atmospheric Pollution: Criteria and Concepts for Operational Modelling of Mercury in Forest and Lake Ecosystems 1

2

2

2,3

M. Meili , K. Bishop , L. Bringmark , K. Johansson , 4 5 6 J. Munthe , H. Sverdrup , W. de Vries 1

Stockholm University, Institute of Applied Environmental Research (ITM), SE-106 91 Stockholm, SWEDEN (E-mail: Markus.Meili@itm.su.se) 2 Swedish University of Agricultural Sciences, Dept. of Environmental Assessment, SE-750 07 Uppsala, SWEDEN 3 Swedish Environmental Protection Agency, SE-106 48 Stockholm, SWEDEN 4 Swedish Environmental Research Institute (IVL), P.O. Box 470 86, SE-402 58 Göteborg, SWEDEN 5 Lund University, Dept. of Chemical Engineering, P.O. Box 124, SE-221 60 Lund, SWEDEN 6 Alterra Green World Research, P.O. Box 47, NL-6700 AA Wageningen, THE NETHERLANDS INTRODUCTION Mercury (Hg) is regarded as a major environmental concern in many regions, because of high concentrations in freshwater fish, and because of potential toxic effects on soil microflora. The main source of Hg in most watersheds is atmospheric deposition, which has increased 2 to >20-fold over the past centuries. A promising approach to support current European efforts to limit transboundary air pollution is the development of emission-exposure-effect relationships, which aim to determine the critical level of atmospheric pollution (CLAP, cf. critical load) causing harm or concern in sensitive elements of the environment. This requires a quantification of slow ecosystem dynamics from short-term collections of data. CONCEPT Here, we present a simple and flexible modelling concept as an operational tool to assess past and future metal contamination of terrestrial and aquatic ecosystems, including ways of minimizing requirements for computation and data collection, focusing on exposure of biota in forest soils and lakes to Hg. Issues related to the complexity of Hg biogeochemistry are addressed by (1) a model design that allows independent validation of each model unit with readily available data, (2) a process- and scale-independent model formulation based on concentration ratios and transfer factors without requiring loads and mass balance, and (3) an equilibration concept that accounts for relevant dynamics in ecosystems without long-term data collection or advanced calculations. MODEL Based on past Swedish data, we present a model to determine CLAP-Hg from standardized values of region- or site-specific synoptic concentrations in four key boreal watershed matrices: precipitation (atmospheric source), large lacustrine fish (aquatic receptor and vector), organic soil layers (terrestrial receptor proxy and temporary reservoir), as well as new and old lake sediments (archives of response dynamics). Key dynamics in watersheds are accounted for by quantifying current states of equilibration in both soils and lakes based on comparison of contamination factors in sediment cores. Future steady state concentrations in soils and fish are determined by projecting survey data. IMPLICATIONS A regional-scale application to southern Sweden suggests that the response of environmental Hg levels to changes in atmospheric Hg pollution is delayed by centuries and initially not proportional

22



among receptors (atmosphere >> soils ≠ sediments > fish; clearwater lakes >> humic lakes). This has implications for the interpretation of common survey data as well as for the implementation of pollution control strategies. Near Hg emission sources, the pollution of organic soils and clearwater lakes deserves attention. Critical receptors, however, even in remote areas, are humic waters, in which biotic Hg levels are naturally high, most likely to increase further, and at high long-term risk of exceeding the current levels of concern: ≤0.5 mg kg-1 w in freshwater fish, and 0.5 mg kg-1w in soil organic matter. If environmental Hg concentrations are to be reduced and kept below these critical limits, virtually no man-made atmospheric Hg emissions can be permitted.

5

continued (as 1985)

-1

Hg in 1-kg pike [mg kg fw]

Hg in 1-kg pike [mg kg-1 fw]

4

3

2 reduced (as 2000)

1 recent

excl. European

natural

tolerable?

0 past

1985

future

Fig. 1. Standardized Hg concentrations in lacustrine fish of the boreal forest zone at various levels of atmospheric pollution over time. Estimates of pre-industrial, recent, potential, and maximum tolerable (critical) values, given as representative regional means for undisturbed watersheds in forested inland regions of southern Sweden (56–59°N). Thin bars show the potential range from precipitation-fed clearwater lakes (low) to runoff-fed polyhumic waters (high); bold bars show the representative range in boreal forest lakes where runoff accounts for 20-80% of the Hg input in the 1980's but 60-95% at steady state. All levels refer to standardized regional means and are based on observations in several hundred watersheds; ranges are larger if accounting for local variability among lakes due to other factors (about fourfold) or among different fish within a lake (about 30-fold). The broken lines show current levels of concern.

REFERENCE Meili, M., Bishop, K., Bringmark, L., Johansson, K., Munthe, J., Sverdrup, H., de Vries, W. (2003) Critical levels of atmospheric pollution: criteria and concepts for operational modelling of mercury in forest and lake ecosystems. Sci. Total Environ. (in press).


23


Formation of Me202Hg from Stable 202Hg Added to the Water Column of a Canadian Shield Lake – Initial Results from the METAALICUS Experiment 1

2

N. Ogrinc , H. Hintelmann 1

Department of Environ. Sci., "J. Stefan" Institute, Jamova 39, 1000 Ljubljana, SLOVENIA (E-mail: nives.ogrinc@ijs.si) 2 Department of Chemistry, Trent University, 1600 W Bank Drive, Peterborough ON K9J 7B8, CANADA (E-mail: hhintelmann@trentu.ca) INTRODUCTION This presentation discusses recent results from the METAALICUS study, a whole-ecosystem stable isotope Hg-addition experiment, carried out at the Experimental Lake Area (ELA) in northwest Ontario, Canada. The aime of the projekt is to evaluate the contribution of atmospherically deposited mercury to the overall levels of methylmercury observed in fish. The study site is the entire watershed of a first order drainage lake (L658), including upland, wetlands, and the lake itself. The full scale Hg addition began in summer 2001 where three different inorganic mercury isotopes were introduced to the lake watershed to examine the relative contributions of these sources to methylation and to MeHg supply to fish. This presentation focuses on water column processes. The lake was spiked with 202 Hg(II) and the kinetic of the partitioning of Hg(II) and the occurance of methylmercury in the dissolved and particulate phase was monitored througout the summer. The distribution of the 202 Hg(II) in the water surface after spiking in June 2001 is shown in Fig. 1. Total 202Hg [ng/l]

Northing

L658

2.2 2.1 2 1.9 1.8 1.7 1.6 1.5 1.4 1.3 1.2 1.1 1 0.9 0.8 0.7

TN

Easting Fig. 1. Spatial distribution of 202Hg(II) in the water surface after first spiking in June 2001.

MATERIAL AND METHODS Water column profiles were collected every two weeks after spiking. Total mercury (HgT) and MeHg in water and in particles were determined using isotope dilution techniques in combination with ICPMS and GC/ICP-MS (Hintelmann and Evans, 1997; Hintelmann and Ogrinc, 2003). With these analytical techniques we were able to track the methylation of Hg isotopes added to the lake.

24



RESULTS AND DISCUSSION The concentrations of 202HgT in dissolved and particulate phases at the surface of the lake showed the expected zig-zag pattern caused by the biweekly additions of 202Hg(II). Furthermore, the 202HgT concentrations in the dissolved phase were high in the epilimnion and the isotope appeared in the water near the sediment surface only later in the season suggesting that most of the Hg was transported on particles down to the bottom. The fraction of a newly formed Me202Hg was larger in the dissolved phase compared to the fraction of Me202Hg found on particles. Levels of Me202Hg continuously increased over the sampling period and eventually contributed more than 25 % to the total MeHg in the hypolimnion (Fig. 2). During the overturn and water column mixing the average concentration of Me202Hg was 0.053 ng/l constituting 33 % of the overall MeHg in the water column.

Me 202Hg [pg/l]

300 250 200 150 100 50

Fig. 2. Distribution of Me202Hg in the water column during the sampling period from June 2001 to March 2002.

CONCLUSIONS The experimental approach of adding stable Hg isotopes to a whole lake allowed us to monitor the production of new MeHg under natural conditions. Seasonal formation rates can be calculated to better predict MeHg bioaccumulation. REFERENCES Hintelmann, H., Evans, R.D. (1997) Application of stable isotopes in environmental studies – Measurements of monomethylmercury (CH3Hg+) by isotope dilution ICP-MS and detection of species transformation, Fresenius J. Anal. Chem., 358, 378-385. Hintelmann, H., Ogrinc, N. (2003) Determination of stable mercury isotopes by ICP/MS and their application in environmental studies, In: Cai, Y., Braids, C.O. (Eds.), Biogeochemistry of environmentally important trace elements, ACS Symp. Ser. Vol. 835, Washington, DC, pp. 321338.


25


Fate of Trace Elements in the Pedosphere of Forest Soils in Alpine Environment, Italy 1

1

1

2

C. Bini , S. Gemignani , L. Zilocchi , F. Corradini , G. Sartori

3

1

Dept of Environmental Sciences, University of Venice. Dorsoduro 2137, Venezia, ITALY (E-mail: bini@unive.it) 2 Istituto Agrario S. Michele all’Adige (Trento), ITALY 3 Museo di Storia Naturale, Trento, ITALY INTRODUCTION Element mobility in the exogenic environment is of major importance with regard to their bioavailability and the potential risk for contamination. Supergenic alteration processes may lead to the release of toxic elements, particularly heavy metals, into the environment. In order to examine the behaviour of such elements, it is necessary to study the interface between rocks, the biosphere and the hydrosphere, that is, the pedosphere. This is a multicomponent complex system whose chemical equilibria are frequently in a thermodynamic steady state. Trace elements of lithogenic and anthropogenic origin occur in soils distributed in various solid -state forms, in the organic fraction and in the soil solution (Alloway, 1992) and their fate in the environment depends upon several factors. In this summary, our objective is: To evaluate the background level of heavy metals in soils of an area with famous environmental sceneries in northern Italy; To ascertain metal mobility and possible contamination of some sites, and the related environmental hazard, with special reference to the pollution of the Dolomites, unique landscapes and delicate ecosystems of the Italian Alps. MATERIAL AND METHODS A regional soil survey was carried out in recent years in forested areas of the Trentino Region, in northern Italy. The Region is characterized prevalently by mountains ranges with steep morphology and different pedolandscapes, related to the nature of the parent material and to the geomorphic surfaces the soils originated from. Limestones and dolostones (approximately 60 %) dominate over crystalline or silicate rocks; glacial and alluvial deposits of mixed lithology are frequent in the lower parts of the valleys. Forestry and pasture are the main land utilization types on these soils (mostly Inceptisols and Mollisols). Approximately 460 soil samples from more than 120 representative soil profiles were sampled and analysed for routine analyses (Soil Survey Manual) and total Cu, Ni, Pb, Cd, Zn, Cr, Fe, Mn contents (AA spectrometry after digestion with HNO3 + HClO4 mixture). RESULTS AND DISCUSSION The territory investigated extends over approximately 400.000 ha (70 % of the whole region), mostly in mountain areas. Beech, spruce and larch are the most important forest components, while limestone is the dominant parent material. The soils examined were grouped in seven Subgroups of the Soil Taxonomy, Inceptisols (50 %) and Mollisols (30 %) being the dominant soil Orders. Most soil horizons (260) have strong acid reaction (mean pH=6.1, range 1.7–8.5), though many of precipitation (mean 1200 mm/y) in the investigated area. The organic matter content is generally high (mean 920 gr/kg) at surface, and decrease with depth (range 1.58-14.60 gr/Kg), the richest horizons being those from calcareous materials (mollic epipedon). The background level of heavy metals (mean total Zn 77 mg/kg; Cu 12 mg/kg; Cr 27 mg/kg; Ni 15 mg/kg; Pb 40 mg/kg; Cd 0,5 mg/kg, Mn 400 mg/kg; Fe 2190 mg/kg) is within the “normal” range for soils of Italy (Angelone and Bini, 1992), with the exception of Zn and Pb

26



concentrations in two mine soils (up to 540 mg/kg and 4300 mg/kg, respectively). Surface horizons, however, present elemental concentrations higher than C horizons (Zn up to 132 mg/kg; Cu 21 mg/kg; Cr 55 mg/kg; Ni 37 mg/kg; Pb 69 mg/kg; Cd 1,4 mg/kg), while mean levels of trace elements in B horizons are lower than in the A and C horizons. This suggests that weathering of parent material releases trace elements, that are complexed by organic ligands in surface horizon. STATISTICAL ANALYSIS. The univariate statistical analysis gave significant positive correlation of some element couples (Ni-Cr 0.96; Zn-Pb 0.89; Cr-Zn 0.73) and soil parameters (C-N 0.95; Ni-clay 0.77; Cr-clay 0.74). The multivariate statistics enabled partition of the 14 variables into 6 principal components that cover globally 90 % of the cumulative variance. Only PC1 (38 % variance) and PC2 (19 % variance) will be discussed here. PC1 comprehends negative scores of trace elements, silt and clay; it may be considered an expression of mineral weathering and mobilization. PC2 presents negative scores of the variables C and N (expression of humification), and positive scores for pH (carbonation). The distribution of variables in the PC1/PC2 space (Fig. 1) allows distinction of strongly weathered horizons from silicate rocks (podzolic soils, group A; leached soils, group B; umbric epipedon, group C) from weakly altered horizons from calcareous rocks (Mollisols, group D), while the less developed soils (Entisols, Inceptisols) are dispersed in the central part of the graph.

Fig. 1. PC1/PC2 score plot. Dots represent mean values of different soil horizons.

CONCLUSIONS Background levels of trace elements in the soils investigated are consistent with “normal” trace element concentration of soils from Western Europe. Therefore, no contamination is recorded in the whole Trentino region. A geological matrix effect may be accounted for metal release by parent material weathering. The high concentration of organic complexing ligants is the major cause of metal accumulation in surface horizons. Statistical analysis enhanced identification of soil groups with different weathering stages, and proved a useful tool to outline pedogenic trends. REFERENCES Alloway, B.J. (1995) Heavy metals in soils (2° edit.), Blackie Academic & Professional, Glasgow, U.K., 362 pp. Angelone, M., Bini, C. (1992) Trace element concentrations in soils and plants of Western Europe. In: Adriano, D. (Ed.), Biogeochemistry of trace metals, Lewis, Boca Raton, 19-60.


27


Geochemical Signature of Contaminated Sediment Remobilization Revealed by Spatially Resolved X-ray Microanalysis of Annual Rings of Salix nigra L. 1,2

3

2

2

1

T. Punshon , A. Lanzirotti , K. McLeod , P.M. Bertsch , J. Burger 1

Consortium for Risk Evaluation with Stakeholder Participation, Environmental and Occupational Health Science Institute, Division of Life Sciences, Rutgers University, 604 Allison Road, Piscataway, NJ 08854, USA 2 Savannah River Ecology Laboratory, University of Georgia, Drawer E, Aiken, SC 29802, USA 3 Consortium for Advanced Radiation Sources, The University of Chicago, 5640 S. Ellis Avenue, Chicago, IL 60637, USA INTRODUCTION Spatially resolved X-ray microanalysis was used to determine the concentration and distribution of metals (specifically Ni and U) within annual rings of willow trees (Salix nigra L.) growing on an eroding former radiological settling pond (Steed Pond: SP), and the impacted riparian depositional area downstream (Tims Branch: TB) on the Savannah River Site. Approximately 44,000 (±26,000) kg of natural and depleted U was released to SP between 1954-1985, in addition to large quantities of Ni, and lesser amounts of Al, Cu, Pb and Cr. Partial failure of the dam occurred in 1984, releasing contaminated sediments into TB, and remained open to erosion during periodic storm events typical of the region. Contamination history and geochemistry are well studied in these areas, making them ideal sites for dendroanalytical studies (Punshon et al., 2002; Sowder et al., 2002). MATERIAL AND METHODS Soil samples were collected from SP, TB and an uncontaminated area (Boggy Gut: BG), dried and extracted with HF and HNO3 to determine total soil metal concentration. Tree core tissue was collected from various willows inhabiting SP, TB and BG. Elemental composition of dried, finely ground ( 5.5), Mn and Zn added in a nutrient solution were immobilized at (extrapolated) rates of 2.1-2.2 and 6.2-40 kg ha-1 yr-1, respectively while in a strongly acid organic layer (pH < 5.5), Zn and Mn were released at 63 and 135 kg ha-1 yr-1, respectively (Wilcke et al., 2002). Thus, at weakly acid sites all micronutrients reaching the soil with throughfall and stemflow may be retained in the organic layer and can therefore not be taken up by the vegetation. As the organic layer is densely rooted, this implies that micronutrients have to be taken up from the atmosphere (as observed) or from the mineral soil where comparatively few roots have been found.

30



During strong rain events, Cu, Mn, and Zn concentrations in stream water increased parallel to water level (Fig. 1) and total organic C concentrations (not shown). At the same time, pH decreased. The reason for this observation was that surface-near acid soil water rich in C and nutrients quickly reached the stream. This was confirmed by the similarity of the δ18O values of the shallow soil water and the stream water during stormflow (own unpublished results). 15

MC1 MC2 MC3

-1

Stream flow [l s ]

a)

10

5

0

b)

Cu Mn Zn

-1

Cu, Mn, Zn [µg l ]

15

10

5

30.04.

09.04.

19.03.

26.02.

05.02.

15.01.

25.12.

04.12.

13.11.

23.10.

02.10.

11.09.

21.08.

31.07.

10.07.

19.06.

29.05.

08.05.

17.04.

27.03.

0

Fig. 1. Temporal course of a) the instantaneous stream discharge of three microcatchments (MC1-3) and b) mean Cu, Mn, and Zn concentrations in stream water.

CONCLUSIONS Canopy uptake and immobilization in laboratory incubation of weakly acid organic layers of Mn and Zn indicate that there may be a lack of micronutrients for plant growth. This lack is probably enhanced by large micronutrient outputs associated with stormflow events when micronutrient concentrations in stream water increase. To test whether micronutrient limitation explains the small tree stature, diagnostic fertilizer experiments are required. REFERENCES Tanner, E.V.J., Vitousek, P.M., Cuevas, E. (1998) Experimental investigation of nutrient limitation of forest growth on wet tropical mountains, Ecol., 78, 10-22. Ulrich, B. (1983) Interactions of forest canopies with atmospheric constituents: SO2, alkali and earth alkali cations and chloride. In: Ulrich, B., Pankrath, J. (Eds.) Effects of Accumulation of Air Pollutants in Forest Ecosystems, D. Reidel Publishing, Dordrecht. pp. 33-45. Wilcke, W., Yasin, S., Abramowski, U., Valarezo, C., Zech, W. (2002) Nutrient storage and turnover in organic layers under tropical montane rain forest in Ecuador, Eur. J. Soil Sci., 53, 15-27.


31


Partitioning of Lead as a Function of Soil Characteristics 1

2

T.V. Pampura , J.E. Groenenberg , K. Terytze

3

1

Institute of Physicochemical and Biological Problems in Soil Science of Russian Academy of Sciences, Pushchino, Moscow region, 142 290, RUSSIA (E-mail: pampura@mail.ru) 2 Alterra, Green World Research, P.O. box 47, Wageningen, 6700 AA, THE NETHERLANDS (E-mail: J.E.Groenengerg@alterra.wag.ur.nl ) 3 Free University, Malteserstr., 74-100, house B, Berlin , 12249, GERMANY (E-mail: konstantin.terytze@uba.de) INTRODUCTION It is recognized now that in many cases the harmful effect of metal on the ecosystem is related to metals dissolved in soil solution, especially uncomplexed free ions. However critical limits and soil quality standards are expressed mostly in total (or “reactive”) metal content in soil. Therefore one has to use an empirical relationship (transfer functions TF) linking HM concentration (or activity) in soil solution with metal content in soil and soil characteristics to calculate a correspondent dissolved metal concentration (and activity). Methods used for investigation of HM partitioning include methods based on the “field” partitioning data, and laboratory batch adsorption experiments. Both have advantages and disadvantages. An analysis of existing methods reveals that they have to be modified to provide data, which can be used for the modeling of metal behavior in the field. Objectives of our study were: (a) To develop a new adsorption technique with soil:solution ratio close to natural conditions. To complement this adsorption method with the investigation of “reactive” pool of metal adsorbed in order to get information on HM partitioning with increasing metal content in terms applicable to the field conditions. (b) To apply this technique to study Pb partitioning in different types of soil in order to derive transfer functions connecting concentration (activity) of metals in soil solution with reactive metal content in soil and soil characteristics. MATERIAL AND METHODS Samples of Niedermoor (10–30 cm), Pararendzina (Loess) (0-30 cm), Braunerde (Of/Oh) contaminated with Pb(NO3)2 (0-500 mg Pb/kg soil) were incubated at 50 % water holding capacity (WHC) at 20°C in darkness for 1 month. Water was then added to give 100 % WHC. After 24 hours soil and solution were separated by centrifugation (5000 rpm). Supernatant pH was measured immediately after centrifuging. Soil solution was analyzed for Pb (GF AAS) and main cations (ICPOES) and anions (IC) including dissolved organic carbon (DOC) (TOC/TN analyzer). The “reactive” metal content in soil was determined by extraction with 1 M NH4NO3 (DIN 19730). Activity of all ions was calculated with the chemical speciation program EPIDIM (Groenendijk, 1995). Complexation of cations with DOC was modelled using a diprotic acid analogue for DOC (Reinds et al., 1995). Multiple linear regression analysis was used to derive transfer function connecting metal concentration (activity) in soil solution, and “reactive” metal content in soil, pH, DOC, CEC effective, organic matter content (equations [1] and [2]): Log[Csolution] = logKC-Q + a·log[SOM] + b·pH + c·log[DOC]+ d· log[CECeffect] + n·log[Qsoil]

[1]

Log[asolution] = logKa-Q + a·log[SOM] + b·pH + c·log[DOC]+ d· log[CECeffect] + n·log[Qsoil]

[2]

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where N- number of data sets (N= 36), Qsoil - reactive metal (NH4NO3 1M extraction), mol/kg, Csolution (asolution ) metal concentration (activity) in soil saturation extract, mmol/l, [SOM] – soil organic matter content, %, pH – pH of soil solution, [DOC] – dissolved organic matter content in soil solution, mol/l, [CECeff] – effective cation exchange capacity of soil, mmol/100 g . The DOC content in mg/l was recalculated in the concentration of organic acids in mol/l using equation [DOC, mol/l] = (DOC, mg/l) · f, where the value of f was set at 5,5 µmolc.mg–1 C (Reinds et al., 1995). RESULTS AND DISCUSSION Results are presented in Table 1. Table contains the values of the equation coefficients corresponding to each independent variable (a, b, c, d, n), equation constant logK, coefficient of determination R2, and the value of the standard error (Y) for the Y (dependent variable). The different rows of table correspond to the different sets of independent variables included in the equation. Table 1. Values for coefficients (equation [1] and [2]).

logKC-Q

a (SOM)

Equation [1] 0.39 -11.55 /3.51/ -22.57 5.09 -23.25 1.39 Equation [2] 4.93 -28.20 8.87 -40.6 10.6 -41.8 4.01

b (pH)

c (DOC)

d (CECeff)

n

R2

(Y)

/1.55/ /1.63/ /-0.14/

/-2.30/ -5.06

3.21

0.76 0.88 0.74 1.03

0.89 0.90 0.91 0.96

0.28 0.28 0.27 0.18

2.92 3.01 -0.16

/-2.59/ -7.52

5.74

1.95 0.85 0.69 1.22

0.77 0.97 0.97 0.99

1.10 0.42 0.41 0.20

/…/ - insignificant for P=0.95

High values of regression coefficients R2 were obtained for the transfer functions for concentration [1] and activity [2] of Pb in soil solution. Applicability of transfer functions were studied by comparison of calculated Pb concentrations with those measured in the field soil saturation extracts for 11 soil samples (Germany), providing big variability in soil characteristics: SOM 1.61 – 76.6 %; pH 3.6-7.8; Clay 0–12.4 %; CEC eff., 18– 317.6 mmol/100 g. In general values predicted with the regressions were in good agreement with field data. Most data demonstrated deviation of less then 5 times from the calculated values. The way to get greater accuracy in prediction of soil solution concentration is to derive individual TF for groups of soils with similar properties. ACKNOWLEDGEMENTS We would like to express our deep gratitude to H.-D. Gregor, R. Hofmann, R. Schleyer (Federal Environmental Agency, Germany), as well as H.-D. Nagel, G. Schütze (Öko-data, Strausberg, Germany), and DAAD, Germany, for organization and support of this work. REFERENCES Reinds, G.J., Bril, J., de Vries, W., Groenenberg, J.E., Breeuwsma, A. (1995) Critical and present loads for cadmium, copper and lead for European forest soils. Wageningen, the Netherlands, DLO Winand Staring Centre, Report 96, 91 pp.


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Long-Term Effects of Sewage Sludge Applications on Cadmium Bioavailability, Distribution and Leaching in Arable Soil P. Bergkvist, N. Jarvis, D. Berggren

Department of Soil Sciences, SLU, P.O. Box 7014, SE-750 07 Uppsala, SWEDEN (E-mail: Petra.Bergkvist@mv.slu.se) INTRODUCTION Use of sewage sludge on arable land may be constrained due to its content of toxic heavy metals, especially the more mobile elements, as cadmium (Cd). Studies have shown that sludge application to arable soils results in an increased bioavailability and crop uptake of cadmium (McGrath et al., 2000). However, there is controversy over the likely long-term effects of sludge-applications on metal uptake, especially with regard to the role of organic matter decomposition following cessation of applications (Chang et al., 1997). The objective of this study was to investigate the long-term effects of sludge applications on cadmium bioavailability and mobility in soil. A field trial, started in 1956 on the effects of organic and inorganic fertilizers on soil chemistry and fertility was utilized for this purpose. MATERIAL AND METHODS The field trial is situated on a clay loam in Ultuna, Uppsala. The measurements were made in 1997 in plots from a treatment that has received sewage sludge biennially for 41 years and a control treatment (calcium nitrate fertilizer). Soil samples from 0-90 cm soil depths were used for measurements on HNO3 and EDTA-extractable Cd, SOC and DOC concentrations, pH and crop Cd concentrations. Batch adsorption experiments on various Cd concentrations and pH values were done to investigate the influence on sludge on Cd solubility and partitioning of Cd between the solid phases and the solution. RESULTS AND DISCUSSION Calculated mass balances showed that 92 % of the accumulated Cd loading was retained in the topsoil after 41 years, while no influence of the Cd loadings was detected deeper than 17 cm below the base of the topsoil, even though enhanced DOC concentrations were found throughout the profile of sewage sludge treated plots. No significant trends in crop Cd concentration with time could be discerned, but compared to the control, the Cd solubility measured in batch experiments was 20 times larger larger in the sludge-amended topsoil, and the straw Cd concentration was also c. 2 times larger. Batch experiments confirmed that soil pH was the dominant factor controlling Cd adsorption at this site, and that organic carbon apparently played a minor role. In contrast, effective kd values for cadmium in the sewage sludge treatment were slightly smaller at any given pH despite much larger organic carbon contents compared to the control (Fig. 1), due to the presence of DOC.

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3

log kd-value, [kg l-1]

Control y = 0,48x - 0,60 R2 = 0,95 p0.05; *=significant at P