Cyanide in Water and Soil: Chemistry, Risk, and Management

CYANIDE in WATER and SOIL

Chemistry, Risk, and Management

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CYANIDE in WATER and SOIL

Chemistry, Risk, and Management David A. Dzombak Rajat S. Ghosh George M. Wong-Chong

Boca Raton London New York

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Preface “Cyanide” is a chemical with a long and fascinating history of respectful and productive use by mankind. The fundamental cyanide species, the cyanide ion CN− , is a highly versatile and strong binder of metals in aqueous solution, a property that has been exploited in ingenious ways for commercial processes that have benefited society. The best known and largest volume uses of cyanide are in the gold mining and electroplating industries. In hydrometallurgical gold mining, aqueous solutions of CN− are used to extract and concentrate gold from ores containing very small amounts of gold. In electroplating, solutions of metal–CN species are used as the baths into which solid metals are dipped and coated with the metal from solution. The deposition of the metal from solution onto the solid metal is governed by the electrochemical gradient induced in the system, and by the metal–cyanide solution chemistry. Cyanide is also produced incidentally in significant quantities in a number of industrial processes, including coal coking and gasification, iron and steel manufacturing, aluminum manufacturing, and petroleum refining. This results in the need for control of cyanide releases in the form of gases, solids, and liquids. The substantial use of cyanide compounds in commerce coupled with the substantial incidental production of cyanide compounds means that significant amounts of cyanide are introduced into the environment on a continuous basis. Cyanide species are frequently occurring contaminants in water and soil. There are also natural sources of cyanide, such as black cherry and cassava plants. Indeed, there is a natural cycle of cyanide. However, anthropogenic inputs of cyanide to the environment are far greater in amount than natural inputs. Of course, “cyanide” is also widely known, and perhaps best known, as a potent human toxin. The most toxic form of cyanide is hydrogen cyanide, HCN, which is as toxic, and often even more so, to wildlife, especially aquatic life. There is great fear of “cyanide” in society, but some chemical forms of cyanide are nontoxic and in fact used regularly in food and cosmetic products. An example is the solid Prussian Blue, or ferric ferrocyanide, which is used as a blue pigment for use in inks, dyes, cosmetics, and other products. The chemistry of cyanide is both complex and diverse, and there are many different chemical forms of cyanide, including solid, gaseous, and aqueous species, and both inorganic and organic species. The particular chemical forms of cyanide that exist in a system, referred to as the speciation of the chemical, are all important in determining the environmental fate, transport, and toxicity of the cyanide. In our careers in environmental engineering and science, we have encountered many different problems involving cyanide in water and soil. Cyanide has been a focus in engineering and research projects that we have performed related to industrial and municipal wastewater treatment, groundwater treatment, industrial waste management, site remediation and restoration, and water quality assessment. These projects have been sponsored by a wide range of companies, industrial research organizations, and regional and federal government agencies. There is widespread interest in cyanide management for environmental and human health protection. We have learned much about cyanide use, management, emissions, and behavior in the environment in the course of these projects. Our education has been aided by useful knowledge and information acquired from many different sources and people. We undertook the preparation of this book to bring together in one place some of the current knowledge and information about cyanide release to, and behavior in, the environment, and means v

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of controlling or remediating these releases. No other broad-based examination of this topic exists. While there has been much good research and engineering development performed in the gold mining industry on cyanide management and control of environmental releases, most notably the work of Dr Terry Mudder and colleagues, this work has been focused on the industry with an orientation toward advancement of hydrometallurgical gold mining. There is much to be learned from the extensive knowledge about cyanide that has been gained in the gold mining industry, but there is a broader range of cyanide challenges in environmental engineering and science. Our book takes on this broader scope. This book tries to address the full range of issues pertaining to cyanide fate, transport, treatment, and toxicity in water and soil. We examine the sources of cyanide released to the environment, both anthropogenic and natural. We have tried to develop an appropriate balance of depth and scope of coverage. There have been compromises made on depth of coverage in some topical areas, but in all areas we have endeavored to provide good and current references to enable the reader to learn more about topics of particular interest. We developed this book to serve as a useful reference tool for engineers and scientists, including both practitioners and researchers, in academia, industrial organizations, government, and engineering and science consulting firms. We hope we have succeeded in our goal. Effective management and remediation approaches for cyanide in the environment require consideration of issues spanning many different fields. In this context, we have collaborated with a wide range of individuals possessing a wide range of expertise in our cyanide-related projects. To address the range of topics that we wanted to examine in this book, we engaged a number of our former and current collaborators to help us with the book. We are most grateful to the contributing authors, listed following this preface and in the header for each chapter. We are also grateful to Alcoa, Inc. and Niagara Mohawk Power Corporation for financial support that helped make this book project possible; and USFilter Corporation, the RETEC Group, Inc. and the Carnegie Mellon University Department of Civil and Environmental Engineering for providing assistance with preparation of graphics and the manuscript. We owe special thanks to Jacqueline Ziemianski, Donna Silverman, and Kacey Ebbitt of the RETEC Group, Inc. for their good work with preparation of graphics and securing permissions for use of copyrighted material, and to Nichole Dwyer of Carnegie Mellon University for her careful work in helping us with revising and formatting the text, with completing and formatting references, and with permissions. Finally, we thank our families for their understanding as we used many hours of family time to work on this book. David A. Dzombak Rajat S. Ghosh George M. Wong-Chong

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Editors

David Dzombak, Ph.D., P.E., DEE, is a professor in the Department of Civil and Environmental Engineering at Carnegie Mellon. Dr Dzombak’s research and professional interests include aquatic chemistry; fate and transport of chemicals in surface and subsurface waters; water and wastewater treatment; in situ and ex situ soil treatment; hazardous waste site remediation; abandoned mine drainage remediation; and river and watershed restoration. He has over 70 peer-reviewed publications and is the joint holder of three patents related to water and soil treatment. He has extensive research and consulting experience with cyanide management and treatment in soils, wastewaters, and process residuals. He has served as a member of the U.S. Environmental Protection Agency Science Advisory Board and is involved with numerous other professional service activities. Dr Dzombak received his Ph.D. in Civil-Environmental Engineering from the Massachusetts Institute of Technology in 1986. He also holds an M.S. in Civil-Environmental Engineering and a B.S. in Civil Engineering from Carnegie Mellon University, and a B.A. in Mathematics from Saint Vincent College. He is a registered Professional Engineer in Pennsylvania, and a Diplomate of the American Academy of Environmental Engineers. Dr Dzombak was elected a Fellow of the American Society of Civil Engineers in 2002. Other awards include the Professional Research Award from the Pennsylvania Water Environment Association (2002); Jack Edward McKee Medal from the Water Environment Federation (2000); Aldo Leopold Leadership Program Fellowship from the Ecological Society of America (2000); Distinguished Service Award from the Association of Environmental Engineering and Science Professors (1999); Walter L. Huber Civil Engineering Research Prize from the American Society of Civil Engineers (1997); Harrison Prescott Eddy Medal from the Water Environment Federation (1993); and National Science Foundation Presidential Young Investigator Award (1991). Rajat S. Ghosh, Ph.D., P.E., is a Program Manager with the EHS Science and Technology Group of Alcoa, Inc., the world’s largest producer of aluminum. He formerly was a Senior Technical Consultant in the Pittsburgh office of The RETEC Group, Inc., a U.S. environmental engineering and consulting company. Dr Ghosh’s research and professional interests are in geochemistry, transport and treatment of inorganic compounds (especially cyanide and heavy metals) in the subsurface; analytical method development for various inorganic and organic compounds; and subsurface multiphase flow and chemistry of organic compounds including coal tar, DNAPLs, and petroleum hydrocarbons. Dr Ghosh has extensive research and consulting experience with the electric power, natural gas, and aluminum industries in the United States in relation to cyanide management and treatment issues in soil and groundwater. In addition, Dr Ghosh serves as a senior technical reviewer for the U.S. Department of Defense basic environmental science and technology development program for site remediation under the auspices of the Strategic Environmental Research and Development Program (SERDP) and Environmental Security and Technology Certification Program (ESTCP). Dr Ghosh received his Ph.D. in Civil-Environmental Engineering from the Carnegie Mellon University in 1998. He also holds an M.S. in Chemical Engineering from University of Wyoming and a B.S. in Chemical Engineering from Jadavpur University, India. He is a registered Professional Engineer in Pennsylvania. He has over 20 professional publications in the open literature and is a joint holder of a U.S. patent on cyanide treatment technology. Dr Ghosh serves as a member of ASTM’s D-19 Committee on Water. Dr Ghosh was elected as a member of the Sigma Xi Honor Society. Other vii

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awards include the Jack Edward McKee Medal from the Water Environment Federation (2000) and the Graduate Student Award from American Chemical Society (1998). George M. Wong-Chong, Ph.D., P.E., DEE, retired director of process wastewater research at USFilter Corporation (Engineering and Construction), has over 35 years of experience in technology development, design, construction, operation, research and teaching of the management and treatment of contaminated groundwater, wastewaters, and solid hazardous waste. Dr Wong-Chong’s experience spans a range of industries including iron and steel, coal tar refining, organic chemicals, petroleum refining, munitions, aluminum manufacturing, coal gasification, live stock agriculture, and municipal wastewater. His experience in the iron and steel industry, where cyanide is a major concern, is internationally recognized; for coke plant wastewaters he developed a patented process, NITE/DENITE™, for the direct biological treatment of flushing liquor, which can contain very high concentrations of ammonia, cyanide, phenols, and thiocyanate. He also holds a patent for the physical/chemical treatment of municipal and industrial wastewaters. Dr Wong-Chong received his Ph.D. in Agricultural Engineering from Cornell University in 1974. He also holds an M.S. in Environmental Engineering from the University of Western Ontario, Canada, and a B.S. in Chemical Engineering from McGill University, Canada. He is a registered Professional Engineer in 10 states, and a Diplomate of the American Academy of Environmental Engineers. In 1999, Dr Wong-Chong received the Pennsylvania Water Environment Association Professional Research Award and the American Institute of Chemical Engineers Pittsburgh Section Award for Outstanding Professional Accomplishments in the Field of Consulting Engineering. Dr Wong-Chong has over 50 publications and presentations to his credit and remains very interested in waste water treatment technology development.

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Contributors

Todd L. Anderson, P.E. Malcolm Pirnie, Inc. Emeryville, CA

Sharon M. Drop, M.S. Alcoa, Inc. Pittsburgh, PA

Barbara D. Beck, Ph.D., DABT DABT, Gradient Corp. Cambridge, MA

David A. Dzombak, Ph.D., P.E., DEE Carnegie Mellon University Pittsburgh, PA

Brice S. Bond, M.S. Southern Illinois University Carbondale, IL

Stephen D. Ebbs, Ph.D. Southern Illinois University Carbondale, IL

Joseph L. Borowitz, Ph.D. Purdue University West Lafayette, IN

Robert W. Gensemer, Ph.D. Parametrix, Inc. Albany, OR

Joseph T. Bushey, Ph.D. Syracuse University Syracuse, NY

Rajat S. Ghosh, Ph.D., P.E. Alcoa, Inc. Pittsburgh, PA

Rick D. Cardwell, Ph.D. Parametrix, Inc. Albany, OR

Cortney J. Higgins, M.S. Carnegie Mellon University Pittsburgh, PA

Jeremy M. Clark Parametrix, Inc. Albany, OR

Gary E. Isom, Ph.D. Purdue University West Lafayette, IN

Rula A. Deeb, Ph.D. Malcolm Pirnie, Inc. Emeryville, CA

Michael C. Kavanaugh, Ph.D., PE, DEE Malcolm Pirnie, Inc. Emeryville, CA

David K. DeForest Parametrix, Inc. Bellevue, WA

Roman P. Lanno, Ph.D. Ohio State University Columbus, OH

Peter J. Drivas, Ph.D. Gradient Corp. Cambridge, MA

Richard G. Luthy, Ph.D., P.E., DEE Stanford University Stanford, CA ix

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Contributors

Johannes C.L. Meeussen, Ph.D. Energy Research Centre of the Netherlands Petten, The Netherlands Charles A. Menzie, Ph.D. Menzie-Cura and Associates Winchester, MA David V. Nakles, Ph.D., P.E. The RETEC Group Pittsburgh, PA Edward F. Neuhauser, Ph.D. Niagara Mohawk Power Co. Syracuse, NY Sujoy B. Roy, Ph.D. Tetra Tech, Inc. Lafayette, CA Mara Seeley, Ph.D., DABT DABT, Gradient Corp. Cambridge, MA Neil S. Shifrin, Ph.D. Gradient Corp. Cambridge, MA

John R. Smith, Ph.D., P.E. Alcoa, Inc. Pittsburgh, PA Angela J. Stenhouse, M.S. Parametrix, Inc. Bellevue, WA Thomas L. Theis, Ph.D., P.E., DEE Univ. of Illinois at Chicago Chicago, IL Jeanne M. VanBriesen, Ph.D. Carnegie Mellon University Pittsburgh, PA George M. Wong-Chong, Ph.D., P.E., DEE USFilter Corporation Pittsburgh, PA Thomas C. Young, Ph.D. Clarkson University Potsdam, NY Anping Zheng, Ph.D. URS Corp. Wayne, NJ Xiuying Zhao, Ph.D. Clarkson University Potsdam, NY

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Contents Chapter 1

Introduction George M. Wong-Chong, David A. Dzombak, and Rajat S. Ghosh

1

Chapter 2

Physical and Chemical Forms of Cyanide Rajat S. Ghosh, David A. Dzombak, and George M. Wong-Chong

15

Chapter 3

Natural Sources of Cyanide George M. Wong-Chong, Rajat S. Ghosh, Joseph T. Bushey, Stephen D. Ebbs, and Edward F. Neuhauser

25

Chapter 4

Manufacture and the Use of Cyanide George M. Wong-Chong, David V. Nakles, and Richard G. Luthy

41

Chapter 5

Physical–Chemical Properties and Reactivity of Cyanide in Water and Soil David A. Dzombak, Rajat S. Ghosh, and Thomas C. Young

57

Chapter 6

Biological Transformation of Cyanide in Water and Soil Stephen D. Ebbs, George M. Wong-Chong, Brice S. Bond, Joseph T. Bushey, and Edward F. Neuhauser

93

Chapter 7

Analysis of Cyanide in Water Rajat S. Ghosh, David A. Dzombak, Sharon M. Drop, and Anping Zheng

123

Chapter 8

Analysis of Cyanide in Solids and Semi-Solids David A. Dzombak, Joseph T. Bushey, Sharon M. Drop, and Rajat S. Ghosh

155

Chapter 9

Fate and Transport of Anthropogenic Cyanide in Surface Water Thomas C. Young, Xiuying Zhao, and Thomas L. Theis

171

Chapter 10

Fate and Transport of Anthropogenic Cyanide in Soil and Groundwater Rajat S. Ghosh, Johannes C.L. Meeussen, David A. Dzombak, and David V. Nakles

191

Chapter 11

Anthropogenic Cyanide in the Marine Environment David A. Dzombak, Sujoy B. Roy, Todd L. Anderson, Michael C. Kavanaugh, and Rula A. Deeb

209

Chapter 12

Cyanide Cycle in Nature Rajat S. Ghosh, Stephen D. Ebbs, Joseph T. Bushey, Edward F. Neuhauser, and George M. Wong-Chong

225

Chapter 13

Human Toxicology of Cyanide Joseph L. Borowitz, Gary E. Isom, and David V. Nakles

237

xi

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Contents

Chapter 14

Aquatic Toxicity of Cyanide Robert W. Gensemer, David K. DeForest, Angela J. Stenhouse, Cortney J. Higgins, and Rick D. Cardwell

251

Chapter 15

Toxicity of Cyanide to Aquatic-Dependent Wildlife Jeremy M. Clark, Rick D. Cardwell, and Robert W. Gensemer

285

Chapter 16

Human Health Risk Assessment of Cyanide in Water and Soil Barbara D. Beck, Mara Seeley, Rajat S. Ghosh, Peter J. Drivas, and Neil S. Shifrin

309

Chapter 17

Ecological Risk Assessment of Cyanide in Water and Soil Roman P. Lanno and Charles A. Menzie

331

Chapter 18

Regulation of Cyanide in Water and Soil David V. Nakles, David A. Dzombak, Rajat S. Ghosh, George M. Wong-Chong, and Thomas L. Theis

351

Chapter 19

Cyanide Treatment Technology: Overview George M. Wong-Chong, Rajat S. Ghosh, and David A. Dzombak

387

Chapter 20

Ambient Temperature Oxidation Technologies for Treatment of Cyanide Rajat S. Ghosh, Thomas L. Theis, John R. Smith, and George M. Wong-Chong

393

Chapter 21

Separation Technologies for Treatment of Cyanide David A. Dzombak, Rajat S. Ghosh, George M. Wong-Chong, and John R. Smith

413

Chapter 22

Thermal and High Temperature Oxidation Technologies for Treatment of Cyanide Rajat S. Ghosh, John R. Smith, and George M. Wong-Chong

439

Chapter 23

Microbiological Technologies for Treatment of Cyanide George M. Wong-Chong and Jeanne M. VanBriesen

459

Chapter 24

Cyanide Phytoremediation Stephen D. Ebbs, Joseph T. Bushey, Brice S. Bond, Rajat S. Ghosh, and David A. Dzombak

479

Chapter 25

Management of Cyanide in Municipal Wastewaters David A. Dzombak, Anping Zheng, Michael C. Kavanaugh, Todd L. Anderson, Rula A. Deeb, and George M. Wong-Chong

501

Chapter 26

Management of Cyanide in Industrial Process Wastewaters George M. Wong-Chong, David V. Nakles, and David A. Dzombak

517

Chapter 27

Cyanide Management in Groundwater and Soil Rajat S. Ghosh, David V. Nakles, David A. Dzombak, and George M. Wong-Chong

571

Index

591

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1 Introduction George M. Wong-Chong, David A. Dzombak, and Rajat S. Ghosh CONTENTS 1.1 1.2 1.3

Cyanide in History . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cyanide Chemical Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cyanide and the Origin of Life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3.1 Role of Hydrogen Cyanide in the Production of Amino Acids . . . . . . . . . . . . . . . . . 1.3.2 Stanley Miller’s Experiment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Ubiquity of Cyanide Compounds in Nature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.1 Cyanide in Outer Space . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4.2 Hydrogen Cyanide in Earth’s Atmosphere. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5 Cyanide in Industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.6 Cyanide Releases to Water and Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.7 Cyanide: Chemistry, Risk, and Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.8 Cyanide Regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.9 Cyanide Treatment Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.10 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2 2 2 2 3 5 5 5 6 6 10 11 11 12 12

Cyanide compounds are produced and used in commerce in large quantities. In the United States, for example, approximately 200 million pounds of sodium cyanide are used annually just in heap leaching extraction of gold from ore [1], with much of this use taking place in one state, Nevada, which accounts for about 70% of U.S. gold production [2]. Large amounts of sodium cyanide are also used in electroplating [3]. Cyanide compounds are also produced incidentally in many processes, such as in aluminum and steel production, and are associated with wastewaters, solid wastes, and air emissions from these processes. In addition, cyanide compounds are present in legacy wastes disposed onsite at numerous manufactured gas plant sites in the United States and Europe. As a result, cyanide is a commonly encountered contaminant in water and soil. Because of the high degree of toxicity in certain forms of cyanide, primarily hydrogen cyanide (HCN), acceptable levels of cyanide compounds in water and soil are generally very low. For example, the U.S. drinking water maximum contaminant level for free cyanide (HCN and CN− ) is 0.2 mg/l, while the U.S. ambient water quality criterion for acute exposures in freshwater systems is 22 µg/l. As this thousandfold difference indicates, some aquatic organisms are significantly more sensitive to cyanide than are humans. Addressing problems of cyanide contamination in water and soil can be very challenging. Complicating factors include the complex chemistry and speciation of cyanide; the analytical challenges of measuring cyanide species in water and soil; the differential toxicity, reactivity, and treatability of the various cyanide species; overlapping and sometimes inconsistent regulations pertaining to cyanide; and the widespread public fear of cyanide, regardless of its form and location. Knowledge in all these areas is needed to develop effective strategies to remedy or manage cyanide 1

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Cyanide in Water and Soil

2

contamination in water and soil. This book presents current scientific understanding and engineering approaches for managing water and soil contamination with cyanide.

1.1 CYANIDE IN HISTORY Cyanide is a chemical well known to the public as a highly toxic agent [4]. For many, the word “cyanide” evokes emotions of death. This perception is prevalent in the history of cyanide dating back to antiquity, long before any understanding of the chemistry of this family of compounds was known. Traitorous Egyptian priests of Memphis and Thebes were poisoned using the pits of peaches [5]. In the 20th century, HCN gas was used in gas chambers in the World War II Holocaust, in prisons for execution of criminals with death sentences, and also as a chemical warfare agent. In 1782, the Swedish chemist Carl Wilhelm Scheele discovered a flammable, water-soluble acidic gas, later identified as HCN, when he heated the cyanide-bearing solid Prussian Blue in an aqueous sulfuric acid solution [6–8]. The name given to the evolved gas was Prussian Blue Acid, also referred to as prussic acid or blue acid [7]. This same gas caused Scheele’s death four years later [8]. The words “cyanine” and “cyanide,” derived from the Greek word “kyanos” for blue, soon came into use to describe the gas [7]. In 1811, Guy Lussac determined the composition of the gas as consisting of one molecule each of carbon, hydrogen, and nitrogen [6]. He referred to the HCN gas as hydrocyanic acid, or hydrogen cyanide.

1.2 CYANIDE CHEMICAL STRUCTURE Cyanide compounds contain the cyano-moiety, which consists of the carbon atom triply bonded to the nitrogen atom (−C≡N). The most basic, and most toxic, of these compounds is hydrogen cyanide (H−C≡N), hydrocyanic acid. HCN is a gas at ambient temperature, and is freely soluble in water. In water, HCN dissociates at high pH (pK a = 9.24 at 25◦ C) to form the cyanide anion, CN− . There are many different inorganic and organic cyanide compounds. Inorganic compounds include simple salts of cyanide with various metals such as sodium cyanide, NaCN(s), potassium cyanide, KCN(s), and more complex solids such as ferric ferrocyanide, Fe4 (Fe(CN)6 )3 (s), also known as Prussian Blue. The simple salts are highly soluble in water. The aqueous solubility of Prussian Blue and other similar complex cyanide solids are functions of pH and redox potential. There are also many organocyanide compounds, such as acetonitrile (CH3 CN), acrylonitrile (CH2 CHCN), and cyanogenic glycosides.

1.3 CYANIDE AND THE ORIGIN OF LIFE 1.3.1 ROLE OF HYDROGEN CYANIDE IN THE PRODUCTION OF AMINO ACIDS In the Precambrian or prebiotic period, about 4.6 billion years ago, primary components of the earth’s atmosphere were carbon monoxide, methane, hydrogen, nitrogen, ammonia, and water [9]. The German biologist E. Pfluger hypothesized that as the earth’s surface slowly cooled from an incandescent mass, HCN was formed by the chemical union of carbon and nitrogen, and that this compound had time to transform and polymerize to form proteins which constitute living matter [10]. Figure 1.1 and Figure 1.2 illustrate the polymerization of HCN, to form adenine, and the reaction of HCN and formaldehyde (another compound formed from the reaction of the constituents in the primitive earth’s atmosphere) to form glycine. These two compounds allow the synthesis of many amino compounds. Oro and Kimball [11,12] demonstrated the synthesis of adenine, a nucleic acid, and other purine intermediates from HCN under possible primitive earth conditions. These abiotically synthesized proteins were important stepping-stones to life, as we know it today.

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Introduction

3

N N

N

N +HCN

+HCN

N Hydrogen cyanide

N (HCN)3

(HCN)2 N N +HCN

N

N

Diaminomaleonitrile (HCN)4

N N N +HCN

N

N

Adenine (HCN)5 Hydrogen

Carbon

N

Nitrogen

FIGURE 1.1 Polymerization of hydrogen cyanide to form adenine. (Source: Barbieri, M., The Organic Codes: An Introduction of Semantic Biology, Cambridge University Press, Cambridge, MA, 2002. With permission.)

1.3.2 STANLEY MILLER’S EXPERIMENT In 1953, Stanley Miller demonstrated that HCN and certain organic compounds, including aldehydes and amino acids, can be formed from the constituents of the prebiotic earth atmosphere, that is, methane, ammonia, hydrogen, and water [9]. The experiments, which earned Miller a Nobel Prize, were performed in a spark-discharge reaction apparatus as shown in Figure 1.3. The apparatus, which was claimed to be a crude model of the primitive earth’s atmosphere, was charged with water and air was evacuated; then, a mixture of ammonia, methane and hydrogen was added. The water in the small flask was boiled to initiate a circulation of gases and water vapor into the reaction flask, in which an electric spark was generated. The spark initiated the reaction of the ammonia, hydrogen,

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Cyanide in Water and Soil

4

+

N

+

N

N N

Formaldehyde

Hydrogen cyanide

Ammonia

Ammonitrile

N N

Ammonitrile

N

+

Water Hydrogen

Carbon

Water

N

+

Ammonia N

+

Glycine Oxygen

Nitrogen

FIGURE 1.2 Reaction of hydrogen cyanide and formaldehyde to form glycine. (Source: Barbieri, M., The Organic Codes: An Introduction of Semantic Biology, Cambridge University Press, Cambridge, MA, 2002. With permission.)

Electrodes

Gases

To a vacuum pump

spark discharge

Water out Condenser Water in Water droplets

Boiling water

Liquid water in trap conaining organic compounds

FIGURE 1.3 Apparatus for experiment by Stanley Miller that demonstrated formation of hydrogen cyanide from constituents of the prebiotic Earth atmosphere. (Source: Miller, S.L. and Orgel, L.E., The Origins of Life on Earth, Prentice-Hall, Englewood Cliffs, NJ, 1974. With permission.)

methane, and water to form HCN and aldehydes. A typical experiment entailed operating of the spark continuously for about 1 week with regular analysis of samples from the system. Figure 1.4 shows the reaction profiles for ammonia (charged material), and amino acids, HCN and aldehydes (reaction products). Miller’s data clearly demonstrated a mechanism for abiotic production of HCN in the atmosphere, one that also exists today during electrical discharges associated with thunderstorms [9].

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5

8

NH3(×10)

Molar concentration

7 Amino acids(×103)

6 5

HCN (×102)

4 3 2

Aldehydes (×103)

1

25

50

75 100 Time (h)

125

150

FIGURE 1.4 Reactant and product concentrations in the experiment by Stanley Miller. (Source: Miller, S.L. and Orgel, L.E., The Origins of Life on the Earth, Prentice-Hall, Englewood Cliffs, NJ, 1974. With permission.)

The findings of Miller are further substantiated by the discovery of the presence of CO, HCN, OH• , formaldehyde and methanol in outer space [13].

1.4 UBIQUITY OF CYANIDE COMPOUNDS IN NATURE Cyanide compounds occur commonly in nature. HCN is present in outer space, in the earth’s atmosphere, in plants, animals, microbes, and fungi. Cyanide can be produced by certain plants, bacteria, fungi, and algae. Chapter 3, which examines natural sources of cyanide, discusses in detail the occurrence, role, and environmental impact of cyanide in plants, animals, microbes, and fungi. The natural cycle of cyanide in the environment is the focus of Chapter 12.

1.4.1 CYANIDE IN OUTER SPACE Hydrogen cyanide has been detected at a number of locations in outer space. For example, it is a trace constituent in the nitrogenous atmosphere of Titan, the largest moon of Saturn [14], and in the coma of the Hale–Bopp comet [15]. Polymerization products of HCN are the dominant components of dust grains sampled from the tail of Comet 81P/Wild2 in 2004 [16]. This presence of HCN in space is now being used to study the birth of massive stars [17]. The detection of large amounts of HCN toward the center of a protostar (an evolving star) means that it has already started to warm up; from this information it is possible to determine the degree of evolution and the age of the star [17].

1.4.2 HYDROGEN CYANIDE IN EARTH’S ATMOSPHERE Hydrogen cyanide is detectable in the troposphere and stratosphere of the earth. Its concentration in the nonurban troposphere of the northern hemisphere has been reported as approximately 160 pptv [18]. In the tropical upper troposphere, a range of HCN concentrations from 200 to 900 pptv

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has been reported [19]. From field measurements and modeling it has been established that biomass burning is a major global source of HCN emissions [19,20]. Estimates of the total release of HCN to the atmosphere from biomass burning range from 1.4 to 2.9 × 1012 g (as N) per year [19]. The residence time of HCN in the atmosphere is approximately two to four months [19]. The oceans of the world provide a sink for the atmospheric releases of HCN and other compounds from biomass burning [19], as discussed in Chapter 11.

1.5 CYANIDE IN INDUSTRY Substantial quantities of cyanide compounds are used and produced in commerce (Chapter 4). Today most cyanide compounds are manufactured starting with HCN, which is synthesized by the platinumcatalyzed reaction of ammonia and methane [3]. HCN is a basic chemical feed stock used in the manufacture of sodium cyanide for gold mining and electroplating; adiponitrile for nylon; methyl methacrylate for clear plastic; triazines for agricultural herbicides; methionine for animal food supplement; and chelating agents (e.g., nitrilotriacetate) for water and wastewater treatment [3]. Worldwide annual production and manufacturing capacity of HCN in 1992 were estimated to be 0.95 million tons and 1.32 million tons, respectively [3]. A 2001 estimate of worldwide cyanide production was 2.60 million tons [7]. In 2001, 0.75 million tons of HCN were produced in the U.S. (Table 1.1). A significant fraction, estimated to range from 8 to 20%, of HCN is used to produce sodium cyanide [3,21,22], much of which is used in hydrometallurgical gold mining. The production and use of cyanide is growing, as indicated by the chronological tabulation of HCN production in the United States in Table 1.1. In addition to use of cyanide compounds in gold mining, electroplating, and chemical production, cyanide compounds are also used in some applications that involve direct distribution to the environment. Sodium ferrocyanide, Na4 (Fe(CN)6 ) and ferric ferrocyanide, Fe4 (Fe(CN)6 )3 (s) are used as an anticaking agent in road salt [23]. It is the presence of ferric ferrocyanide that gives a blue color to salt in which it is used. These compounds can dissolve in water after placement on road surfaces. Sodium ferrocyanide is also used in some forest fire retardants [24].

1.6 CYANIDE RELEASES TO WATER AND SOIL Most cyanide that occurs in water and soil is anthropogenic, derived from industrial processes, but there are natural sources of cyanide as noted above. The combination of widespread industrial sources and natural sources leads to detectable concentrations of cyanide in many natural waters, though concentrations are usually low. In a 1981 evaluation of monitoring data in the USEPA STORET database, it was determined that the mean concentration of total cyanide in surface waters of the United States did not exceed 3.5 µg/l, but in 37 of 50 states there were sampling locations where total cyanide concentrations in excess of this level were reported [25]. Sample results from a number of industrialized areas had total cyanide concentrations greater than 200 µg/l. Total cyanide concentrations in U.S. drinking water intake supplies are usually very low (<10 µg/l) to nondetectable [26,27]. Analyses of six Canadian surface waters for a performance comparison of analytical techniques yielded total cyanide concentrations less than 12 µg/l for the three streams sampled, but in the range of 19 to 49 µg/l for the three lakes sampled (Table 1.2). The major sources of cyanide in water and soil are discharges and wastes from metal mining processes, metal manufacturing and finishing processes, chemical production, and petroleum refining [4,28,29]. Cyanide contamination of water and soil from industrial sources can be recent, from ongoing operations, or may be from wastes disposed long ago. For example, cyanide occurs in soil and groundwater in the vicinity of old spent potlining landfills at aluminum smelting facilities [30,31], and in old disposal areas for oxide box wastes at former manufactured gas plant sites [31–33].

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TABLE 1.1 Production of Hydrogen Cyanide in the United States, 1983–2001 Year

Production, 103 tons/yr

2001 2000 1999 1998 1997 1996 1995 1994 1993 1992 1991 1990 1989 1988 1987 1986 1985 1984 1983

750 765 745 725 710 695 675 645 600 570 565 585 565 500 470 430 365 365 330

Sources: Production estimates for 1983–1988: Data from Pesce, L.D., Kirk-Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. Production estimates for 1989–2001: Data from Myers, E., American Chemistry Council, Washington, DC, personal communication, 2002.

TABLE 1.2 Concentrations of Free Cyanide and Total Cyanide in Six Surface Water Samples from Across Canada Free cyanide (µg/l)

Total cyanide (µg/l)

Sample

Electrode

Colorimetry

Electrode

Colorimetry

Stream 1 Stream 2 Stream 3 Lake 1 Lake 2 Lake 3

4 6 4 5 10 17

3 4 4 6 12 19

7 10 11 21 25 48

8 12 12 19 27 49

Source: Reprinted from Water Res., 104, Sekerka, I. and Lechner, J.F. Potentiometric determination of low levels of simple and total cyanides, 479, copyright (1976), with permission from Elsevier.

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TABLE 1.3 Examples of Discharges from Gold Mine Heap Leaching Operations Location Baia Mare, Romania

Release date/period

Release scenario

Reference

January 30–February 2, 2000

100,000 m3 (26 million gallons) of

[34]

Gold Quarry Mine, Nevada, USA

June 6, 1997

Omai, Guyana

August 19–24, 1995

USMX Mine, Utah, USA

March 11–14, 1995

Summitville, Colorado, USA

1986–1992

cyanide-bearing tailings released due to tailings dam failure 245,000 gallons cyanide solution leakage from heap leach pad; discharge to two nearby creeks 4.2 million m3 (1.1 billion gallons) of cyanide-bearing tailings water released due to tailings dam failure 7 million gallons treated leach solution containing 0.2 ppm cyanide, released from storage ponds to East Fork of Beaver Dam Wash Sustained cyanide solution leaks from heap leach pad, from transfer pipes, and from tailings pond; discharge to Alamosa River

[44]

[45]

[46]

[47,48]

The most dramatic releases of cyanide to water and soil have occurred in the failure of or substantial leakage from heap leaching pads or tailings ponds associated with gold mining operations. Table 1.3 lists some large-volume discharges that have occurred since 1992. The failure of the tailings pond dam at a gold mine near Baia Mare, Romania, in January 2000 provides an example of the large scale of impact that can result from such discharges. Due to heavy precipitation coupled with a rapid snowmelt, a gold mine tailings pond near Baia Mare filled to capacity and overflowed, resulting in washout of a section of the earthen containment dam for the pond. Approximately 100,000 cubic meters of tailings water containing free cyanide, metal–cyanide complexes, metals, and suspended solids were discharged from January 30 to February 2, 2000 [34]. Based on cyanide concentrations in the tailings pond and the approximate spill volume, it is estimated that 50 to 100 tons of cyanide were released. As shown in the map on Figure 1.5, the spill entered the Sasar River, which subsequently joins with the Lapus River, and then the Somes River. The Somes flows into Hungary, and there it discharges into the Tisza River, which flows through Hungary and into Serbia (formerly, Yugoslavia). Just north of Belgrade the Tisza discharges into the Danube, which returns to Romania and eventually discharges into the Black Sea. It took the plume of contamination about 14 days to reach the Danube, which is approximately 800 km in river distance from the spill location. The plume then traveled an additional 1,200 km in the Danube. Total cyanide concentrations as high as 32.6 mg/l were measured in the Somes River at the Hungarian–Romanian border. The maximum cyanide concentration observed in the Tisza River at the Hungarian–Yugoslavian border was 1.5 mg/l, and in the Danube River near the Yugoslavian–Romanian border was 0.34 mg/l. These concentrations, while demonstrating the dilution, biodegradation, and volatilization of the cyanide during riverine transport to the Black Sea, nevertheless were 15 to 1,500 times greater than water quality criteria to protect freshwater aquatic life to acute exposures. As a result, massive fish kills were experienced due to the cyanide plume from Baia Mare (Figure 1.6). An estimate of dead fish in the Hungarian portion of the Tisza River as a result of the spill was 1,240 tons [34]. There were also substantial but unquantified fish kills in the Tisza River in Yugoslavia. Smaller in scale but more widespread are the many continuing releases of cyanide from solid wastes disposed on land in the past and from ongoing wastewater discharges. Cyanide-bearing oxide

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UKRAINE

SLOVAKIA Miskoic 1

2

Baia Mare

za

Budapest

Ti

HUNGARY

4

3 MOLDOVA

Szeged 5

ROMANIA Timisoara

6 Belgrade

7

9

8

Bucharest sar

Sa

YUGOSLAVIA

Black sea

BUICARIA

FIGURE 1.5 Map showing the river transport route for the cyanide plume from the spill at Baia Mare, Romania, January–February 2000. (Source: Data from: UNEP, http://www.rec.org/REC/Publications/ CyanideSpill/ENGCyanide.pdf, 2000.)

FIGURE 1.6 Worker removing dead fish killed by a cyanide spill in Hungary’s Tisza River, at Kiskore on February 9, 2000. Photo by Laszlo Balogh. © Reuters 2000. Used with permission.

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box wastes at thousands of former manufactured gas plant (MGP) sites throughout the United States and Europe are an example of a widely distributed industrial legacy waste. These wastes, which contain the iron cyanide solid Prussian Blue, were disposed in onsite landfills at many MGP sites. Dissolved cyanide is generated by contact of these solids with groundwater, resulting in underground plumes of contamination that can move significant distances, depending on subsurface conditions [31,35,36]. At a former MGP site in Wisconsin, it was demonstrated that dissolved cyanide moved with the groundwater through the sand and gravel aquifer beneath the site, toward a municipal drinking water supply well located 500 m from the site [31]. In another area of Wisconsin, oxide box wastes from an MGP operation were placed as landfill material in three-foot thick layers along an electric transmission line corridor, amounting in just one section of the corridor to 26,000 tons of fill material [37]. Remediation efforts involving removal of the material commenced in the 1990s. Related legal actions eventually resulted in settlements totaling $21.8 million against the responsible company [38,39]. Thus, even localized cyanide contamination problems can have significant technical, regulatory, and legal implications.

1.7 CYANIDE: CHEMISTRY, RISK, AND MANAGEMENT The management and regulation of cyanide in water and soil can be very challenging because of the complexity of the chemistry and toxicology of cyanide, the risk it poses in different environmental contexts, and stringent regulatory requirements to be satisfied [32,40,41]. Many different chemical forms of cyanide occur in water and soil, including dissolved free cyanide (HCN, CN− ), metal4− cyanide complexes (e.g., Ni(CN)2− 4 , Fe(CN)6 ), and organocyanide (e.g., acetonitrile, CH3 CN) species, as well as metal-cyanide solids (e.g., ferric ferrocyanide, Fe4 (Fe(CN)6 )3 (s)). In addition, HCN in water can volatilize, forming HCN(g). The different chemical forms of cyanide and their reactivity and properties are discussed in Chapters 2, 5, and 6. Each of these species is formed and affected by different chemical reactions, and each has different physical, chemical, and toxicological properties. For example, the toxicological significance of each individual metal–cyanide complex is determined by its ability to release free cyanide (CN− or HCN), the target species of concern, under pertinent exposure conditions. The chemical dissociative properties of each complex thus control the release of free cyanide and hence toxicity. Thus, the differences in properties mean that the various cyanide species vary in their toxicity to animals and plants, in their fate and transport in the environment (Chapters 9–11), and in their treatability by physical, chemical, and biological treatment technologies (Chapters 19–24). Until recently, regulation and management of cyanide in water and soil have been focused on total (inorganic) cyanide content (Chapter 18). This focus has been driven in large part by the availability of a long-standing, simple, robust technique for measuring total inorganic cyanide content: strong acid digestion to transform all inorganic cyanide compounds to HCN followed by distillation to volatilize and capture the HCN(g). While such total cyanide measurements are useful and indeed continue to be the predominant means of monitoring cyanide, they provide no direct information about cyanide speciation. The focus on total cyanide content has made regulations of cyanide in water and soil confusing and inconsistent, and has led to management and treatment approaches of varying effectiveness. Knowledge of cyanide speciation is critical to technically and economically effective management of cyanide in water and soil. This is now fairly well recognized in the engineering, science, and regulatory professional communities, but measurements, regulations, treatment technologies, and site management plans with a species-specific focus are still evolving for cyanide. We are in the midst of transitioning to species-specific approaches with respect to cyanide, similar to the transition that occurred through the 1980s for management of metal contaminants in water and soil. This book is intended to help with and to accelerate that transition.

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1.8 CYANIDE REGULATIONS Cyanide aqueous discharge regulations in the United States are based on effluent discharge limitations for categorical industries (e.g., iron and steel, organic chemicals manufacturing, and electroplating) and receiving water quality criteria. Effluent limits for categorical industries are largely based on the performance of Best Available Technology Economically Achievable (BAT) and are contained in the U.S. Code of Federal Regulations (40CFR, Parts 425–471). Thirteen major industry categories are identified in the federal regulations, including 43 subcategories with cyanide discharge limits (see Chapter 18). It is interesting to note that the gold mining industry, a major cyanide user and discharger, is not included in these industry categories. Receiving water quality criteria, which are based on aquatic toxicity studies and aimed at protection of aquatic life, are significantly more stringent than BAT discharge limits. For discharges to water bodies in which the designated use mandates protection of aquatic life, cyanide effluent limits are usually developed with the objective of not exceeding water quality criteria. Discharge limits based on consideration of water quality criteria tend to be very stringent, and can be at or below detection limits achieved in routine commercial analyses. For example, effluent limits for shallow-water marine discharges in the United States are often set at the marine water quality criterion of 1 µg/l free cyanide. The detection limit for free cyanide with standard analytical methods often exceeds this amount by factors of 2 to 5 or more. Chapters 7 and 8 discuss the analytical issues associated with measuring cyanide in water, wastewater, soil and sludges, and the very troublesome issues of detection, practical quantitation limits, and measurement precision. In addition, water quality criteria reflect toxicity to very sensitive aquatic species that may not be present in a particular receiving water. Soil cleanup standards for cyanide have been established by some states in the United States and by some countries in Europe (see Chapter 18). Many other government organizations have established soil screening or action levels to define when additional remedial investigation or action is needed. Soil cleanup standards or screening levels for cyanide vary widely. For example, soil cleanup standards for free cyanide in residential surface soils, where direct human contact can occur, have been set at 30, 160, 1,600, and 4,400 mg/kg by Florida, Maryland, New Jersey, and Pennsylvania, respectively. Free cyanide cleanup standards for nonresidential surface soils established by the same four states are 39,000, 4,100, 56,000, and 23,000 mg/kg. By contrast, the Netherlands has set the “intervention value” for free cyanide in soil at the low value of 20 mg/kg based on human health risk considerations, and has established separate values for complexed cyanide at 50 mg/kg for soils with pH ≥ 5 and at 650 mg/kg for soils with pH < 5 [42]. The Dutch soil “target values” for protection of ecosystems are even lower: 1.0 mg/kg for free cyanide and 5.0 mg/kg for complexed cyanide [42]. Acceptable concentrations of free and complexed cyanide in water and soil are determined by risk assessment. Chapters 13 to 17 examine the toxicity and risk issues that drive the establishment of cyanide aqueous discharge limits and treatment/management objectives for soil and other cyanidecontaminated media.

1.9 CYANIDE TREATMENT TECHNOLOGY An array of technologies is available for the treatment of cyanide in surface water and groundwater, wastewaters, and contaminated soils and sludges. These technologies, discussed in detail in Chapters 19–24, span the gamut of biological, chemical, electrolytic, physical, and thermal treatment processing. Example applications of the technologies employed most commonly in municipal and industrial settings are presented in Chapters 25–27. An important message from these examinations of cyanide contamination management is that commercial applications of the technologies in an economical mode of operation may not yield treated water, soil, or sludge with cyanide concentrations that meet specified regulatory limits. Careful evaluation of technology performance, including treatability testing, is needed prior to application of technologies for cyanide management.

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1.10 SUMMARY AND CONCLUSIONS • Cyanide compounds are produced and used in commerce in large quantities. • Many different chemical forms of cyanide can exist in water and soil, each of which has different physical, chemical, and toxicological properties. • Because of the high degree of toxicity of certain forms of cyanide, primarily hydrogen cyanide (HCN), acceptable levels of cyanide compounds in water and soil can be very low, for example, 1 µg/l for free cyanide in marine waters of the United States. • Many aquatic organisms are significantly more sensitive to cyanide than are humans. • Cyanide species can be formed in nature by both abiotic and biotic processes. Cyanide can be produced by certain plants, bacteria, fungi, and algae. • Background concentrations of cyanide in water and soil are very low. Most cyanide found in water and soil is the result of anthropogenic contamination from industrial sources. • The major sources of cyanide in water and soil are discharges and wastes from metal mining processes, metal manufacturing and finishing processes, chemical production, coal conversion processes, and petroleum refining. • The management and regulation of cyanide in water and soil can be very challenging because of the complexity of the chemistry and toxicology of cyanide and, accordingly, the risk it poses in different environmental contexts. A further complication is that there is widespread public fear of cyanide, regardless of its form and location. • The focus on total cyanide content has made regulations of cyanide in water and soil confusing and inconsistent, and has led to management and treatment approaches of varying effectiveness. • We are in the midst of transitioning to species-specific approaches with respect to cyanide, similar to the transition that occurred for management of metal contaminants in water and soil.

REFERENCES 1. MPC, The last American dinosaur: The 1872 Mining Law, Mineral Policy Center, Washington, DC, http://www.earthworksaction.org/ewa/pubs/MPCfs_LastAmericanDinosaur.pdf, 2004. 2. Dobra, J.L., The U.S. Gold Industry 2001, Special Publication 32, Nevada Bureau of Mines and Geology, Reno, NV, http://www.nbmg.unr.edu/doxftp.htm, 2002. 3. Pesce, L.D., Cyanides, in Kirk-Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. 4. ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997. 5. Lechtenberg, M. and Nahrstedt, A., Naturally occurring glycosides, in Cyanogenic Glycosides, Ikan, R., Ed., John Wiley & Sons, Inc., Chichester, UK, 1999, Chapter 5. 6. Bunce, N. and Hunt, J., History of cyanide, University of Guelph, Department of Physics, http://www.physics.uoguelph.ca/summer/scor/articles/scor176.htm, 2004. 7. Young, C.A., Cyanide: just the facts, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals & Materials Society, Warrendale, PA, 2001, p. 97. 8. Scheele, C.W., The Chemical Essays, Translated from the Transactions of the Academy of Sciences, Academy of Sciences, Stockholm, London, 1786. 9. Miller, S.L. and Orgel, L.E., The Origins of Life on the Earth, Prentice-Hall, Englewood Cliffs, NJ, 1974. 10. Oparin, A.I., The Origin of Life, Dover Publications, Mineola, NY, 1953. 11. Oro, J. and Kimball, A.P., Synthesis of purines under possible primitive earth conditions 1. Adenine from hydrogen cyanide, Arch. Biochem. Biophys., 94, 217, 1961. 12. Oro, J. and Kimball, A.P., Synthesis of purines under possible primitive earth conditions. 2. Purine intermediates from hydrogen cyanide, Arch. Biochem. Biophys., 97, 292, 1962.

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13. Murphy, S., The interstellar medium: Larger molecules in space, Bristol University, School of Chemistry, UK, http://www.chm.bris.ac.uk/webprojects2002/, accessed: June 11, 2004. 14. NASA, Titan, National Aeronautics and Space Administration, Jet Propulsion Laboratory, http://voyager.jpl.nasa.gov/science/saturn_titan.html, accessed: June 11, 2004. 15. Jewitt, D., Hale–Bopp: Hydrogen cyanide, University of Hawaii, http://www.ifa.hawaii.edu/faculty/ jewitt/submm_hb_hcn.html, accessed: June 11, 2004. 16. Kissell, J., Krueger, F.R., Silen, J., and Clark, B.C., The cometary and interstellar dust analyzer at Comet 81P/Wild2, Science, 304, 1774, 2004. 17. ESA, Toxic compound in space signals starbirth, European Space Agency, Science Programme, http://sci.esa.int/science-e/www/object/index.cfm?fobjectid=28635, accessed: June 11, 2004. 18. Cicerone, R.J. and Zellner, R., The atmospheric chemistry of hydrogen cyanide, J. Geophys. Res., 88, 10689, 1983. 19. Li, Q., Jacob, D.J., Bey, I., Yantosca, R.M., Zhao, Y., Kondo, Y., and Notholt, J., Atmospheric hydrogen cyanide (HCN): Biomass burning source, ocean sink? Geophys. Res. Lett., 27, 357, 2000. 20. Holzinger, R., Warneke, C., Jordan, A., Hansel, A., Lindinger, W., Sharffe, D.H., Schade, G., and Crutzen, P.J., Biomass burning as a source of formaldehyde, acetaldehyde, methanol, acetone, acrylonitrile and hydrogen cyanide, Geophys. Res. Lett., 26, 1161, 1999. 21. CMR, Chemical profiles: Hydrogen cyanide, Chemical Market Reporter, http://www.the-innovationgroup.com/ChemProfiles/Hydrogen%20Cyanide.htm, accessed: June 11, 2004. 22. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books Ltd., London, 1991. 23. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 24. Milstein, M., Forest service tells fire retardant maker to remove cyanide, The Oregonian, September 20, 2000. 25. Fiksel, J., Cooper, C., Eschenroeder, A., Goyer, M., and Perwak, J., Exposure and risk assessment for cyanide, EPA-440/4-85-008, U.S. Environmental Protection Agency, Washington, DC, 1981. 26. McCabe, L.J., Symons, J.M., Lee, R.D., and Robeck, G.G., Survey of community water supply systems, J. Am. Water Works Assoc., 62, 670, 1970. 27. Towill, L.E., Drury, J.S., Whitfield, B.L., Lewis, E.B., Galyan, E.L., and Hammons, A.S., Review of the environmental effects of pollutants. V. Cyanide, EPA-600/1-78-027, U.S. Environmental Protection Agency, Cincinnati, OH, 1978. 28. Baker, D.C. and Chou, C.C., Cyanide occurrence and treatment in the petrochemical industry, in Conference on Cyanide and the Environment, Tucson, AZ, Colorado State University, Fort Collins, CO, 1984, p. 379. 29. Lordi, D.T., Lue-Hing, C., Whitebloom, S.W., Kelada, N., and Dennison, S., Cyanide problems in municipal wastewater treatment plants, J. Water Pollut. Control Fed., 52, 597, 1980. 30. Blayden, L.C., Hohman, S.C., and Robuck, S.J., Spent potliner leaching and leachate treatment, in Proc. Light Metals 1987, Denver, CO, The Minerals, Metals and Materials Society, Warrendale, PA, 1987, p. 663. 31. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 32. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 33. Shifrin, N.S., Beck, B.D., Gauthier, T.D., Chapnick, S.D. and Goodman, G., Chemistry, toxicology and human health risk of cyanide compounds in soils at former manufactured gas plant sites, Regul. Toxicol. Pharmacol., 23, 106, 1996. 34. UNEP, Spill of liquid and suspended waste at the Aurul S.A. Retreatment Plant in Baia Mare, Romania, United Nations Environment Programme, Geneva, Switzerland, http://www.uneptie.org/pc/mining/library/publications/assmnt.htm, 2000. 35. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H. and de Haan, F.A.M., Dissolution behavior of iron cyanide (Prussian Blue) in contaminated soils, Environ. Sci. Technol., 26, 1832, 1992. 36. Meeussen, J.L., van Riemsdijk, W.H. and van der Zee, S.E.A.T.M., Transport of complexed cyanide in soil, Geoderma, 67, 73, 1995. 37. Van de Kamp Nohl, M., Power failure, Milwaukee Magazine, January, 2001, p. 56.

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38. Hawkins, L., Wisconsin Energy settles cyanide suit, Milwaukee Journal Sentinel, September 4, 2002. 39. Johnson, A. and Held, T., Deal reached over tainted wood chips, Milwaukee Journal Sentinel, May 28, 2002. 40. Deeb, R.A., Dzombak, D.A., Theis, T.L., Ellgas, W. and Kavanaugh, M.C., The cyanide challenge, Water Environ. Technol., 15, 35, 2003. 41. Nakles, D.V., MGPs and risk assessment, Soil & Groundwater Cleanup, June, 1998, p. 4. 42. Swartjes, F.A., Risk-based assessment of soil and groundwater quality in the Netherlands: Standards and remediation urgency, Risk Anal., 19, 1235, 1999. 43. Myers, E., American Chemistry Council, Washington, DC, personal communication, 2002. 44. LVRJ, Company still seeks cause of cyanide spill, Las Vegas Review Journal, June 18, 1997. 45. Beebe, R.R., Process considerations before and after failure of the Omai Tailings Dam, August 10 to 24, 1995, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 3. 46. NDEP, Hazardous Materials Report, Office of Emergency Management Report, H-950313E, Nevada Division of Environmental Protection, Bureau of Mining Regulation and Reclamation, Carson City, NV, 1995. 47. Bigelow, R.C., Plumlee, G.S., and Edelman, P., The Summitville Mine and its downstream effects, Open File Report 95-23, U.S. Geological Survey, Denver, CO, http://pubs.usgs.gov/of/1995/ofr-9523/summit.htm, 1995. 48. USEPA, Summitville mine Superfund site, U.S. Environmental Protection Agency, Region 8, http://www.epa.gov/region8/superfund/sites/co/sville.htm, accessed: June 8, 2004.

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and Chemical Forms of 2 Physical Cyanide Rajat S. Ghosh, David A. Dzombak, and George M. Wong-Chong CONTENTS 2.1 2.2

Gaseous Forms of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Aqueous Forms of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.1 Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2 Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2.1 Weak Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2.2 Strong Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.3 Cyanate and Thiocyanate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.4 Organocyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3 Solid Forms of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.1 Simple Metal–Cyanide Solids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2 Metal–Metal Cyanide Solids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2.1 Alkali/Alkaline Earth Metal–Metal Solids . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2.2 Other Metal–Metal Cyanide Complex Salts . . . . . . . . . . . . . . . . . . . . . . . . . 2.4 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

17 17 17 19 19 19 19 20 20 21 21 21 22 22 23

Cyanide occurs in many different forms in water and soil systems. The specific form of cyanide determines the environmental fate and transport of cyanide, as well as its toxicity. Understanding the specific form(s) of cyanide present in a particular water, soil, or sediment is critical for assessment of how to manage or treat the cyanide present. This cannot be overemphasized! While “cyanide” is often discussed as a single entity in the popular press and even in professional publications, this is a misleading portrayal. The various forms of cyanide are quite different in their reactivity and their toxicity. Proper professional evaluation, assessment, and design activities pertaining to cyanide contamination management requires knowledge about and careful consideration of cyanide speciation. This chapter provides an introductory overview of the various forms of cyanide that can exist in water and soil systems. All of the remaining chapters of this book assume a basic knowledge of the speciation of cyanide as presented here. A detailed examination of the properties and reactivity of the most commonly occurring aqueous, gaseous, and solid forms of cyanide is provided in Chapter 5. In water and soil systems, cyanide occurs in various physical forms, including many different kinds of species dissolved in water, many different solid species, and several gaseous species. The cyanide species that occur in the aqueous, solid, and gas phases are indicated in Figure 2.1. Chemically, cyanide can be classified into inorganic and organic forms, as indicated in Figure 2.1. Inorganic forms, which occur in all three physical states, include free cyanide, weak metal–cyanide complexes, strong metal–cyanide complexes, thiocyanate and metal–thiocyanate complexes, cyanate 15

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WATER

Free cyanide

Metal–cyanide complexes

Cyanate, thiocyanate

Organocyanides

HCN, CN –

Weak complexes: Ag(CN)2–, CdCN–, … Strong complexes: Fe(CN)4– Fe(CN)3– 6,… 6,

CNO–, SCN–

Nitriles, cyanohydrins, …

GAS

Free cyanide HCN(g)

SOLID

Cyanogen halides CNCl(g), CNBr(g)

Simple metal cyanide solids NaCN(s), KCN(s), CuCN(s) …

Alkali or alkaline earth metal-metal cyanide solids K3Fe(CN)6(s), K4Fe(CN)6(s), KAg(CN)2(s), …

Other metal-metal cyanide solids Fe4[Fe(CN)6 ]3(s), Fe3[Fe(CN)6]2(s), …

FIGURE 2.1 Forms and species of cyanide in water and soil.

and metal–cyanate complexes, and cyanogen halides. Aqueous free cyanide is the sum of hydrogen cyanide, HCN, and its deprotonated form, the cyanide anion, CN− . HCN is volatile under environmental conditions and occurs as both aqueous and gaseous species. Many metals can bond with the cyanide anion to form dissolved metal–cyanide complexes, as well as metal–cyanide solids. Cyanate, CNO− , requires the presence of strong oxidizing agents for its formation and thus is rarely found in the environment. Thiocyanate, SCN− , can be formed in the environment and is also present in a variety of industrial wastewater discharges. The cyanogen halides of interest, CNCl and CNBr, form upon chlorination or bromination of water containing free cyanide. These species are volatile under environmental conditions, and thus occur as both aqueous and gaseous species. Organic cyanides contain carbon–carbon covalent bonding between hydrocarbon and cyanide moieties, and are usually present as dissolved species. Natural as well as anthropogenic sources discharge a wide range of cyanide species to the environment. Over 2650 species of plants (130 families) produce cyanogenic glycosides as part of natural defense mechanisms (Chapter 3). Upon stress or injury, cyanogenic glycosides are hydrolyzed by a coexisting plant enzyme and release HCN. In addition, almost all fruit-bearing plants release HCN during ethylene synthesis, which aids in the fruit ripening process (Chapter 3). Cyanide (as free, organic and metal-complexed cyanide compounds) is used as a raw material during the production of chemicals (nylon and plastics), pesticides, rodenticides, gold, wine, anticaking agents for road salt, fire retardants, cosmetics, pharmaceuticals, painting inks, and other materials (Chapter 4). Cyanide is also used directly in a variety of processes, including electroplating and hydrometallurgical gold extraction (Chapter 4). One of the earliest uses of cyanide dates back to 1704, when the solid phase iron–cyanide compound ferric ferrocyanide (FFC), Fe4 [Fe(CN)6 ]3 (s), also referred to as Prussian Blue, was first used as a pigment for artist colors [1,2]. In addition, free cyanide, weak and strong metal–cyanide complexes, and thiocyanates also occur as by-products of many current and former industrial processes (Chapter 4). Current industries that produce cyanide as a by-product include chemical manufacturing, iron and steel making, petroleum refining, and aluminum smelting. An example of a past industry that generated cyanide-bearing wastewaters and solid wastes in substantial quantities is gas manufacture by coal gasification. There are thousands of former manufactured gas plant (MGP) sites throughout the eastern and midwestern United States and

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Europe with soil containing FFC, which was generated as a process by-product and often managed onsite as fill [3]. Cyanide contamination exists at many other former industrial sites. It is one of the most common contaminants identified at Superfund sites in the United States [4]. The aim of this chapter is to provide an overview of the common physical and chemical forms of cyanide that occur in water and soil systems. In the following sections, the cyanide species of primary interest in gaseous form, dissolved in water, and in solid form are listed and briefly described.

2.1 GASEOUS FORMS OF CYANIDE Three gaseous forms of cyanide are of interest in water and soil systems: hydrogen cyanide (HCN), cyanogen chloride (CNCl), and cyanogen bromide (CNBr). Cyanogen chloride and cyanogen bromide are disinfection by-products formed in water and wastewater treatment [5,6]. HCN is present in wastewater discharges and leachates from certain industrial waste sites, and can be formed in nature as well. Hydrogen cyanide gas is colorless with an odor of bitter almonds. It is highly toxic to humans (see Chapter 13). HCN(g) is very soluble in water, forming a weak acid, HCN(aq), upon dissolution. HCN has a high vapor pressure (630 mm Hg at 20◦ C; Ref [7]) and is readily volatilized from water at pH values less than 9, where HCN remains fully protonated. The cyanogen halides CNCl and CNBr are also colorless gases with high vapor pressures (1230 mm Hg and 121 mm Hg at 25◦ C for CNCl and CNBr, respectively [8,9]). Like hydrogen cyanide gas, CNCl and CNBr are highly toxic to humans if inhaled or absorbed. These are soluble in water, but degrade by hydrolysis, very rapidly at high pH [5]. Degradation is rapid at any pH if there is free chlorine or sulfite present [5]. At pH 10, degradation of CNCl and CNBr by hydrolysis occurs with half-lives in the range of 20 to 40 min [5]. The hydrolysis degradation product is cyanate ion (CNO− ), which can subsequently hydrolyze to CO2 and NH3 at alkaline pH conditions (see Chapter 5).

2.2 AQUEOUS FORMS OF CYANIDE Common aqueous forms of cyanide, listed in Table 2.1, can be broadly divided into four major classes: free cyanide, metal–cyanide complexes, cyanate and thiocyanate species, and organocyanide compounds. Free cyanide comprises molecular HCN and cyanide anion. Metal–cyanide complexes range from weak metal–cyanide complexes (e.g., complexes of copper, zinc, and nickel with CN− ) to strong metal–cyanide complexes (e.g., complexes of cobalt and iron with CN− ). Cyanate and thiocyanate form by oxidation of free cyanide, in the presence of sulfide compounds in the case of thiocyanate. Both of these species are anionic for the environmental pH range, and form complexes with metals. Finally, there are organocyanide complexes, where the cyanide anion is covalently bonded to a hydrocarbon group.

2.2.1 FREE CYANIDE Free cyanide represents the most toxic cyanide forms (see Chapters 13 and 14). It refers to either soluble hydrogen cyanide, HCN(aq), or soluble cyanide anion (CN− ). HCN(aq) is a weak acid with a pKa of 9.24 at 25◦ (Chapter 5). It can dissociate into cyanide ion according to the following dissociation reaction: HCN(aq) = H+ + CN− ,

pK a = 9.24 at 25◦ C

(2.1)

where the “=” sign denotes a two-way, equilibrium reaction. Thus, at pH values less than 9.24, HCN is the dominant free cyanide species, while at greater pH values cyanide ion dominates free cyanide.

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TABLE 2.1 Common Aqueous Cyanide Species Classification

Cyanide species

Free cyanide Weak metal–cyanide

HCN, CN−

2− − AgCN(OH)− , Ag(CN)− 2 , Ag(CN)3 , Ag(OCN)2 2− CdCN− , Cd(CN)02 , Cd(CN)− 3 , Cd(CN)4

complexes

2− 3− Cu(CN)− 2 , Cu(CN)3 , Cu(CN)4

2− − + 0 Ni(CN)02 , Ni(CN)− 3 , Ni(CN)4 , NiH(CN)4 , NiH2 (CN)4 , NiH3 (CN)4 2− Zn(CN)02 , Zn(CN)− 3 , Zn(CN)4

2− − 2− 2− HgCN+ , Hg(CN)02 , Hg(CN)− 3 , Hg(CN)4 , Hg(CN)2 Cl , Hg(CN)3 Cl , Hg(CN)3 Br

Strong metal–cyanide

− BaFe(CN)2− 6 , BaFe(CN)6

− 2− 0 CaFe(CN)2− 6 , CaFe(CN)6 , Ca2 Fe(CN)6 , CaHFe(CN)6

complexes

3− 2− 0 Fe(CN)4− 6 , HFe(CN)6 , H2 Fe(CN)6 , Fe2 (CN)6

K2 H2 Fe(CN)06 , K3 HFe(CN)06 , KHFe(CN)2− 6 3− K2 Fe(CN)2− 6 , KFe(CN)6 2− 2− LiFe(CN)3− 6 , Li2 Fe(CN)6 , LiHFe(CN)6

Fe(CN)3− 6

2− MgFe(CN)− 6 , MgFe(CN)6 2− 2− NH4 Fe(CN)3− 6 , (NH4 )2 Fe(CN)6 , NH5 Fe(CN)6 2− 2− NaFe(CN)3− 6 , Na2 Fe(CN)6 , NaHFe(CN)6

SrFe(CN)− 6

TlFe(CN)3− 6 Au(CN)− 2

Co(CN)3− 6 Pt(CN)2− 4 Cyanate

HOCN, OCN−

Metal–cyanate complexes

Ag(OCN)− 2 , and others

Thiocyanate

HSCN, SCN−

Metal–thiocyanate

MgSCN+ MnSCN+

complexes

FeSCN+ − 0 + FeSCN2+ , Fe(SCN)+ 2 , Fe(SCN)3 , Fe(SCN)4 , FeOHSCN

CoSCN+ , Co(SCN)02 CuSCN+ , Cu(SCN)02 NiSCN+ , Ni(SCN)02

CrSCN2+ , Cr(SCN)+ 2

2− CdSCN+ , Cd(SCN)02 , Cd(SCN)− 3 , Cd(SCN)4

2− ZnSCN+ , Zn(SCN)02 , Zn(SCN)− 3 , Zn(SCN)4 , and others

Organocyanides

Nitriles (e.g., acetonitrile) Cyanohydrins Cyanocobalamin and others

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2.2.2 METAL–CYANIDE COMPLEXES The cyanide anion is a versatile ligand that reacts with many metal cations to form metal–cyanide complexes. These species, which are typically anionic, have a general formula of M(CN)n− x , where M is a metal cation, x is the number of cyanide groups, and n is the ionic charge of the metal–cyanide complex. The stability of metal–cyanide complexes is variable and requires moderate to highly acidic pH conditions in order to dissociate. Metal–cyanide complex dissociation yields free cyanide: + − M(CN)n− x = M + xCN

(2.2)

Metal–cyanide complexes are classified into two broad categories, namely, weak metal–cyanide complexes and strong metal–cyanide complexes, based on the strength of the bonding between the metal and the cyanide ion. Complexes with greater strength of the metal–cyanide bond are more stable in aqueous solution, that is, they dissociate only to a limited extent, and the dissolution process may be very slow.

2.2.2.1 Weak Metal–Cyanide Complexes Weak metal–cyanide complexes are those in which the cyanide ions are weakly bonded to the metal cation, such that they can dissociate under mildly acidic conditions (pH = 4 to 6) to produce free cyanide. Because of their dissociative nature, they are often regulated along with free cyanide in water. Common examples of weak metal–cyanide complexes include copper cyanide (Cu(CN)2− 3 ), 2− 2− ), nickel cyanide (Ni(CN) ), cadmium cyanide (Cd(CN) ), mercury zinc cyanide (Zn(CN)2− 4 4 4 cyanide (Hg(CN)2 ), and silver cyanide (Ag(CN)− 2 ).

2.2.2.2 Strong Metal–Cyanide Complexes Strong metal–cyanide complexes include cyanide complexes with transition heavy metals such as, iron, cobalt, platinum, and gold that require strong acidic conditions (pH < 2) in order to dissociate and form free cyanide. Strong metal–cyanide complexes are much more stable in aqueous solution than the weak ones and are relatively less toxic. Common examples of strong metal–cyanide com3− − plexes include ferrocyanide (Fe(CN)4− 6 ), ferricyanide (Fe(CN)6 ), gold cyanide (Au(CN)2 ), cobalt 2− cyanide (Co(CN)3− 6 ), and platinum cyanide (Pt(CN)4 ).

2.2.3 CYANATE AND THIOCYANATE Free cyanide can be oxidized to form cyanate, CNO− , or, depending on the pH, its protonated form HOCN (pK a = 3.45 at 25◦ C). Cyanate is substantially less toxic than free cyanide. It is rarely encountered in aqueous systems, as a strong oxidizing agent and a catalyst are required for conversion of free cyanide to CNO− or HOCN [10]. When cyanate does form it can react with metals to form metal–cyanate complexes, though these reactions have not been studied extensively (Chapter 5). Free cyanide can react with various forms of sulfur to form thiocyanate, SCN− , which is relatively nontoxic. The two forms of sulfur in the environment most reactive with free CN− are polysulfides, 0 Sx S2− , and thiosulfate, S2 O2− 3 (Chapter 5). Thiocyanate can protonate to form HCNS , but this rarely occurs in natural systems as the pKa for this reaction is 1.1. Thiocyanate can form complexes with many metals (Chapter 5).

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H,R (sugar

O)n

C

C⬅N

R

FIGURE 2.2 General structure of cyanogenic glycosides (R represents CH3 group). Amygdalin (Cherry, Apricot) HO

Dhurrin (Cassava)

CH2OH

CN O

HO

CH2OH O

HO HO

OCH2

C

CH3

O CN

HO

O

OH

OCH

HO

CH3

OH OH

FIGURE 2.3 Common plant cyanogenic glycosides.

2.2.4 ORGANOCYANIDE COMPLEXES Organic cyanide compounds contain a cyanide functional group that is attached to a carbon atom of the organic molecule via covalent bonding. Common examples include nitriles, such as acetonitrile (CH3 CN) or cyanobenzene (C6 H5 CN), which are used as industrial solvents and as raw materials for making nylon products and pesticides. Nitriles can also exist in the natural environment in shale oils [11], in plants [12], or as a plant-growth hormone [13]. Several classes of nitriles can be produced naturally or synthesized chemically, the most common of which are the cyanogenic glycosides and cyanohydrins. Cyanohydrins, also known as α-hydroxynitriles, are organic cyanides with the general structure R1 R2 C(OH)(CN), where the hydroxide group and the cyanide group are attached to the same carbon atom. Cyanogenic glycosides are produced by the plants under natural environmental conditions to aid in their defense mechanism (Chapter 3). These species comprise a cyanide anion that is covalently bonded to a carbon atom, which in turn is bound by a glycosidic linkage to one or more sugars depicted in Figure 2.2. Some common cyanogenic glycosides produced by plants are shown in Figure 2.3. Certain groups of nitriles such as, cyanogenic glycosides, exhibit high stability in water as far as dissociation to free cyanide is concerned. Other organocyanide compounds of interest include cyanocobalamin, also known as Vitamin B12 . It consists of single cyanide group bonded to a central trivalent cobalt cation. Vitamin B12 is synthesized by microorganisms, not by plants, and is found in animal tissues as a result of intestinal synthesis [14]. It is essential for human life, serving numerous functions and being an especially important vitamin for maintaining healthy nerve cells and aiding the production of genetic building blocks DNA and RNA [15]. There are cyanide and noncyanide forms of Vitamin B12 . The noncyanide forms include methylcobalamin, adenosylcobalamin, chlorocobalamin, and hydroxycobalamin. These compounds, also produced by microorganisms, are less stable than cyanocobalamin but also essential to human life.

2.3 SOLID FORMS OF CYANIDE In systems with metals and cyanide present in sufficient quantities, metals can react with cyanide to form a wide range of solids. The solid forms of cyanide may be divided into two general

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TABLE 2.2 Common Solid Phase Cyanide Species Classification

Cyanide species

Simple metal–cyanide solids

KCN(s) NaCN(s) AgCN(s) CuCN(s) Hg(CN)2 (s)

Alkali or alkaline earth metal–metal cyanide solids

K4 Fe(CN)6 (s) K3 Fe(CN)6 (s) K4 Ni4 (Fe(CN)6 )3 (s) K2 CdFe(CN)6 (s) K2 Cu2 Fe(CN)6 (s) KZn1.5 Fe(CN)6 (s)

Other metal–metal cyanide solids

Fe4 [Fe(CN)6 ]3 (s) Fe3 [Fe(CN)6 ]2 (s) Fe[Fe(CN)6 ](s) Fe2 [Fe(CN)6 ](s) Ag4 Fe(CN)6 (s) Cd2 Fe(CN)6 (s) Cu2 Fe(CN)6 (s) Zn2 Fe(CN)6 (s)

categories: simple metal–cyanide solids, which are relatively soluble, and metal–metal cyanide complex solids with varying degree of solubility. Some common metal–cyanide and metal–metal cyanide solids are listed in Table 2.2.

2.3.1 SIMPLE METAL–CYANIDE SOLIDS This class of cyanide solids consist of structurally simple, metal cyanides of the form M(CN)x , where M is an alkali, alkaline earth metal or a heavy metal. Common examples include sodium cyanide (NaCN(s)), potassium cyanide (KCN(s)), calcium cyanide, (Ca(CN)2 (s)), zinc cyanide (Zn(CN)2 (s)), and others (see Table 2.2). Most of these solids are highly soluble in water and readily dissociate, releasing the cyanide ion, and therefore are potentially toxic.

2.3.2 METAL–METAL CYANIDE SOLIDS This class of cyanide solids consists of one or more alkali, alkaline earth, or transition metal cations combined with an anionic metal–cyanide complex. Based on whether the metal cation is alkali/alkaline earth or transition metal, this class of compounds is again subdivided into two categories: alkali/alkaline earth metal–metal cyanide solids and other metal–metal cyanide solids. In the latter, the metals involved are B-type or transition metals [16]. 2.3.2.1 Alkali/Alkaline Earth Metal–Metal Solids This class of structurally complex solids comprises one or more alkali or alkaline earth metal cations ionically bonded to an anionic metal–cyanide complex with the general formula of Ax [M(CN)y ] · nH2 O, where A is an alkali or alkaline earth metal cation (or ammonium ion), M is a transition metal atom, x is the number of alkali metal atoms, y is the number of cyanide groups,

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and n is the number of water molecules incorporated in the solid structure. A common example of this class of compound is potassium ferrocyanide (K4 Fe(CN)6 (s)). Alkali/alkaline earth metal– metal cyanide complex salts can readily dissociate in aqueous solutions, releasing the alkali metal cation and the anionic metal cyanide complex according to the following equation: Ax [M(CN)y ] · nH2 O = xA+ + [M(CN)y ]m−

(2.3)

where m is the ionic charge of the metal–cyanide complex released to solution. 2.3.2.2 Other Metal–Metal Cyanide Complex Salts This class of structurally complex compound comprises one or more transition metal cations ionically bonded to an anionic transition metal cyanide complex with the general formula of Mx [M(CN)y ]z · nH2 O where M is a B-type or transition metal cation, x number of transition metal cations, y is the number of cyanide groups, z is the number of metal–cyanide complexes, and n is the number of water molecules in the structure. Due to the versatility of the cyanide anion as a ligand, there are many different kinds of metal–metal cyanide compounds that exhibit a wide range of structural properties [17]. Metal–metal cyanide solids involving all B-type and transition metals are very stable and relatively insoluble under acidic and neutral conditions (Chapter 5). However, under alkaline conditions, these compounds are relatively soluble, releasing metal cations and anionic metal–cyanide complexes to solution according to the following general reaction: Mx [M(CN)y ]z · nH2 O = xM+ + z[M(CN)y ]m−

(2.4)

where m is the ionic charge of the metal–cyanide complex released to aqueous solution. A well-known example of a transition metal–metal cyanide is ferric ferrocyanide Fe4 (Fe(CN)6 )3 (s), or Prussian Blue, which has various commercial and medicinal uses (Chapter 4).

2.4 SUMMARY AND CONCLUSIONS • Cyanide is present in gas, liquid, and solid forms in water and soil systems. • Many different species of cyanide occur in water and soil systems. The specific form of cyanide determines the environmental fate and transport of cyanide, as well as its toxicity. Understanding the specific form(s) of cyanide present in a particular water, soil, or sediment is critical for assessment of how to manage or treat the cyanide present. • Cyanide mostly occurs in inorganic forms. The dissolved forms of primary interest are free cyanide (HCN and CN− ) and metal–cyanide complexes. Solid forms of cyanide include simple metal–cyanide solids (e.g., NaCN(s), KCN(s)), which are relatively soluble, and more complex, less soluble metal–metal cyanide solids (e.g., Fe4 (Fe(CN)6 )3 (s), or Prussian Blue). The gaseous form of cyanide of primary interest is HCN(g). • Free cyanide, either in dissolved (HCN and CN− ) or gaseous form (HCN(g)), are the species of primary interest with respect to human health and aquatic toxicity. • Dissolved inorganic metal–cyanide complexes can be categorized as weak metal–cyanide complexes and strong metal–cyanide complexes, based on the strength of the bonding between the metal and the cyanide ion. • Cyanate (CNO− ) is formed from oxidation of free cyanide. It can react with metals and form metal–cyanate complexes. • Thiocyanate (SCN− ) is formed from reaction of free cyanide with various forms of sulfur. It can react with metals to form metal-thiocyanate complexes.

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• Organic compounds containing cyanide are produced by both natural and anthropogenic activities. They consist of molecules with carbon–carbon covalent bonding with the –CN group. Common organocyanide compounds include the nitriles, such as acetonitrile (CH3 CN).

REFERENCES 1. ACC, The Chemistry of the Ferrocyanides, American Cyanamid Co., New York, NY, 1953. 2. Feller, R.L., Ed., Artist’s Pigments: A Handbook of Their History and Characteristics, National Gallery of Art, Washington, DC, 1986. 3. Hayes, T.D., Linz, D.G., Nakles, D.V., and Leuschner, A.P., Eds., Management of Manufactured Gas Plant Sites, Vol. 1 & 2, Amherst Scientific Publishers, Amherst, MA, 1996. 4. USEPA, Common chemicals found at Superfund sites, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, http://www.epa.gov/superfund/resources/chemicals.htm, accessed: March 22, 2005. 5. Xie, Y. and Hwang, C.J., Cyanogen chloride and cyangen bromide analysis in drinking water, in Encyclopedia of Analytical Chemistry, Meyers, R.A., Ed., John Wiley & Sons, Chichester, UK, 2000, p. 2333. 6. Zheng, A., Dzombak, D.A., and Luthy, R.G., Formation of free cyanide and cyanogen chloride from chlorination of POTW secondary effluent: laboratory study with model compounds, Water Environ. Res., 76, 113, 2004. 7. ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997. 8. CDC, NIOSH emergency response card: Cyanogen chloride, Centers for Disease Control and Prevention, http://www.bt.cdc.gov/agent/cyanide/erc506-77-4.asp, accessed: April 3, 2005. 9. IPCS/INCHEM, Cyanogen bromide, International Programme on Chemical Safety and the Commission of the European Communities, http://www.inchem.org/documents/icsc/icsc/eics0136.htm, accessed: April 3, 2005. 10. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books Ltd., London, 1991. 11. Evans, E.J., Batts, B.D., Cant, N.W., and Smith, J.W., The origin and significance of nitriles in oil shale, Org. Geochem., 8, 367, 1985. 12. Knowles, C.J., Microorganisms and cyanide, Bacteriol. Rev., 40, 652, 1976. 13. Stowe, B.B. and Hudson, V.W., Growth promotion in pea stem sections. III. By alkyl nitriles, alkyl acetylenes and insect juvenile hormones, Plant Physiol., 44, 1051, 1969. 14. Gershoff, S.N., Vitamin B12, AccessScience@McGraw-Hill, http://www.accessscience.com, accessed: April 3, 2005. 15. UMD, Vitamin B12 (Cobalamin), University of Maryland Medical Center, http://www.umn.edu/altmed/ ConsSupplements/VitaminB12Cobalamincs.html, accessed: April 3, 2005. 16. Stumm, W. and Morgan, J.M., Aquatic Chemistry, Wiley-Interscience, New York, 1996. 17. Dunbar, K.R. and Heintz, R.A., Chemistry of transition metal cyanide compounds: modern perspectives, Prog. Inorg. Chem., 45, 283, 1997.

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3 Natural Sources of Cyanide George M. Wong-Chong, Rajat S. Ghosh, Joseph T. Bushey, Stephen D. Ebbs, and Edward F. Neuhauser CONTENTS 3.1

Cyanide in Vascular Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.1 Cyanogenesis in Plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.2 Role of Cyanogenesis in Plants and its Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.2.1 Impact of Cyanogenesis in Cassava (Manihot esculenta) . . . . . . . . . . . 3.1.2.2 Impact of Cyanogenesis in Forage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.3 Ethylene Production in Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Cyanide in Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.1 Cyanogenic Algae . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2 Cyanogenic Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.3 Cyanogenic Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.4 Role of Cyanogenic Compounds in Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3 Cyanide in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.1 Cyanogenic Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.2 Biosynthesis of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3.3 Role of Cyanogenic Compounds in Arthropods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4 Cyanide in Forest Fires . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

26 26 27 29 30 32 32 32 34 34 35 36 36 36 37 37 38 38

If one considers the beginning of organic life, it is not necessary to pay attention first of all to carbon dioxide and ammonia, because both represent the end and not the beginning of life. The beginning is to be found to a very much larger degree in cyanogens (CN) (E. Pfluger, 1875) [1]

The four most common atoms that make up life on earth are carbon, hydrogen, nitrogen, and oxygen, and hydrogen cyanide (HCN) is comprised of three of these essential macronutrients. Cyanide played a pivotal role in the prebiotic development of amino acids, peptides, nucleotides, lipids, and membranes, and continues to be an integral part of nature on earth and the universe. Cyanide is a source of nitrogen, a nutrient that all living organisms require. Cyanide has been shown to form nutrient “microcycles” in the environment [2,3] (discussed in detail in Chapters 6 and 12). These microcycles involve both cyanogenic (cyanide producing) organisms, as well as organisms that assimilate cyanide as a sources of carbon and nitrogen for growth as discussed in Chapter 6. Today, cyanogenic compounds (compounds containing the CN− moiety) can be found in more than 3000 species of plants, animals, microbes, and fungi. In most organisms, cyanide is used to deter herbivory or pathogenic attack and to regulate specific biochemical processes. The purpose of this chapter is to present an overview of the various natural sources of cyanogenic compounds that are produced by plants, microorganisms (algae, bacteria, fungi), and animals. The plant kingdom appears to have the greatest diversity of species capable of producing significant 25

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quantities of free cyanide and cyanogenic compounds. For each of these organismal groups, the discussion in this chapter will cover the following topics: • • • •

List of organisms that produce cyanogenic compounds Forms of the cyanogenic compounds produced Reaction pathways involved in the synthesis and release of HCN Impact of HCN release on biological systems and its role in the survival of the organism

In addition, the chapter presents a discussion on forest fires as a natural source of cyanide.

3.1 CYANIDE IN VASCULAR PLANTS All higher plants produce HCN in small amounts as part of normal metabolic process, such as ethylene synthesis and nitrile metabolism [4]. Some 2650 plant species, belonging to about 550 genera and more than 130 families, also produce significant quantities of cyanogenic glycosides, a group of polar, water-soluble compounds stored in the plant cell vacuoles [4]. Table 3.1 presents a partial list of cyanogenic plants, primarily those that are economically significant food crops (e.g., cassava, corn, and lima beans), forages (e.g., alfalfa, sorghum, and sudan grasses), and horticultural plants (e.g., ornamental cherry and laurel). Additional discussion of some cyanogenic plants, including cassava (Manihot esculenta) and forages, is presented later in this chapter. Table 3.2 lists the reported cyanide content in various plant tissues. The information in Table 3.2 depicts the following trends with respect to the cyanogenic characteristics in various plants: • Plants of different families contain different quantities of cyanogenic material. • Plants within the same family, and even within the same species, contain different quantities of cyanogenic material (e.g., Acacia spp.); some may be totally acyanogenic [5]. • Different tissues (including whole roots, root cortex, root peels, pods, and leaves) of a plant can contain different concentrations of cyanogenic compounds. • Variations in the HCN or cyanogenic glycoside content within various species may be due to a host of factors, including (a) cultivation condition; (b) soil quality; (c) life stage of the plant when sampled; (d) physiological status of the plant; (e) sampling procedures adopted to sample various tissues, and (f) analytical procedures used to extract cyanide from various plant tissues. • The cyanogenic content of the plants listed can reach levels considered to be toxic to humans and animals. Based on available information, an acute dose of HCN for humans can vary from 0.5 to 3.5 mg kg−1 of body weight [6].

3.1.1 CYANOGENESIS IN PLANTS Cyanogenesis in plants usually refers to the enzymatic liberation of HCN from cyanogenic glycosides. This occurs when plant tissue is damaged or stressed, or when the cyanogenic glycosides in seeds are released to provide the embryo with a source of nitrogen. The release of HCN was first detected scientifically in 1802 [4]. The cyanogenic plant tissue generally contains both cyanogenic glycoside and lipid, and hydrolytic enzymes (e.g., β-glucosidase and hydroxynitrile lyase), stored in separate compartments. When the glycosides and the decomposing enzymes come together, the result is hydrolysis of the cyanogenic glycoside and the liberation of HCN. Figure 3.1 illustrates the decomposition of the cyanogenic glycoside, linamarin, by linamarase and hydroxylnitrile lyase. The general result of these enzymatic reactions is the release of HCN, along with a sugar or fatty acid, and an aldehyde or ketone; cyanogenic lipids degrade with the release of a fatty acid. Some 91 cyanogenic glycosides and four cyanolipids of plant origin have been identified and their structures elucidated [4,7]. Table 3.3 presents a partial list of cyanogenic glycosides, plant sources,

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TABLE 3.1 A List of Common Cyanogenic Plants Scientific classification Acacia spp. Bahia oppositifolia Bambusa spp. Cercocarpus spp. Eucalyptus cladocalyx Florestina tripteris Glyceria striata Heteromeles arbutifolia Holcus lanatus Hydrangea spp. Lathyrus spp. Leguminosae spp. Linaria striata Linum spp. Lotus corniculatus Lupinus spp. Manihot esculenta Phaseolus lunatus Prunus spp. Pyrus malus Rosaceae spp. Sorghum spp. Stillingia suckleyana Trifolium repens Triglochin spp. Vicia sativa Zea mays

Common name Catclaw, Whitehorn Bahia Bamboo Mountain mahogany Eucalyptus Florestina Fowl mannagrass Christmas berry Velvet grass Hydrangea Beans, peas, vetch Alfalfa legumes Herb Flax Birdsfoot trefoil Lupines Cassava Lima bean Apricot, cherry, peaches, plum Apple Cherry laurel Sudan grass, Johnson grass Poison suckleya White clover Arrowgrass Vetch seed Maize (corn)

Sources: Data from Jones, D.A., Cyanide Compounds in Biology, Evered, D. and Garnety, S.F., Eds., John Wiley & Sons, 1988, p. 151; Bolhuis, G.G., Netherlands J. Agr. Sci., 2, 176, 1954; Seigler, D.S., Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 133.

and degradation products; this table shows the glycosides are all degraded to produce HCN, an aldehyde, and glucose. A complete list of 91 cyanogenic glycosides with their plant sources and structure is presented by Lechtenberg and Nahrstedt [4]. The biosynthesis of cyanogenic glycosides generally starts with one of the following amino acids: L-valine, L-isoleucine, L-leucine, L-phenylalanine, and L-tyrosine; other precursor materials identified are nicotinic acid and 2-2 -cyclopentenyl glycine [4]. Cyanogenic lipids are all derived from leucine, with the glucose moiety replaced by a fatty acid of varying chain length and degree of unsaturation [7]. Figure 3.2 presents a generalized biosynthesis pathway for cyanogenic glycoside an example pathway for the synthesis of dhurrin from L-tyrosine.

3.1.2 ROLE OF CYANOGENESIS IN PLANTS AND ITS IMPACTS Cyanogenesis in plants is generally recognized as a chemical defense mechanism against herbivory and pathogenic attack. However, co-evolution has allowed some animals to circumvent this

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TABLE 3.2 Cyanide Content in the Tissue of Selected Plants

Plant species Cassava (bitter) Casava (bitter) Cassava (bitter) Cassava (bitter) Cassava Cassava (sweet) Cassava (sweet) Bamboo Sorghum

Tissue analyzed Whole root Root pulp Root peel Dried root cortex Whole root Leaves Leaves (fresh) Root pulp Root peel Whole root Leaves Immature shoot tip Whole immature plant Whole immature plant Dried pulp

Sorghum forage Lima beans (Java) (Puerto Rico) (Burma) Lima beans Bitter Almonds Acacia erioloba Acacia hebeclada Acacia sieberiana var. woodii Acacia cheelii Acacia conninghamii Acacia doratoxylon Acacia glaucescence Eucalyptus polyanthemos Eucalyptus yarraensis

Leaves Unripe pod Leaves Leaves Unripe pod Unripe pod Leaves Leaves Leaves Leaves Leaves Leaves

Cyanide concentration (mgHCN kg FW−1 ) 530 310 650 2450 395 310 80–4000 38 200 462 468 8000 2500 108 249 100–800 Occasional >1000 3120 3000 2000 100–3120 2500 1059–1807 882 1924 1394–2317 1100 1134 1237 118–236 234 2513 0–181 39–113

Reference [50] [51] [52] [52] [53] [51] [52] [52] [52]

[54] [52]

[50] [50] [52] [52] [52]

[52] [52] [52] [52] [55] [55]

defense. For example, the larvae of the southern armyworm (Spodoptera eridania Cramer) showed a preference for the foliage of the cyanogenic lima bean [5]; army ant larvae feed on the cyanogenic grass (Cynodon plectostachyus) [8] and fungal leaf blight on the cyanogenic rubber tree (Hevea brasiliensis) [9]. Another example involves the cyanogenic plant birdsfoot trefoil (Lotus corniculatus), larvae of the five spotted burned moth (Zygaena trifolii), snails (Helix aspersa), and an ichneumonid wasp (Apantales zygaebarun) [10]. Production of cyanide by birdsfoot trefoil deters herbivory by the snail species, with the snail feeding only on tissues with low cyanide concentrations. In contrast, the moth larvae preferentially consume highly cyanogenic material and sequester the cyanide, to be used as their own defense compound. This deters predation, except by A. zygaebarum, which can detoxify the cyanide present in the larvae internally.

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CH2OH

29

CH2OH

CN

O O

C CH3

HO

+

CN

O OH

CH3 H2O

b-glucosidase

OH

+ HO

HO

OH

C

CH3

CH3 Acetone cyanohydrin

OH

OH b-D-glucopyranose

Linamarin

CN hydroxynitrile HO

C

CH3

CH3

HCN

+

O

lyase

Acetone cyanohydrin

C

CH3

CH3 Hydrogen cyanide

Acetone

FIGURE 3.1 Decomposition of linamarin by plant enzymes. (Source: Adapted from Conn, E.E., Toxicants Occurring Naturally in Foods, National Academy of Science, Washington, DC, Chapter 14, 1973. With permission.)

The impacts of cyanogenesis in plants are discussed in relation to the cultivation of cassava and forage crops, sorghum, and sudan grass.

3.1.2.1 Impact of Cyanogenesis in Cassava (Manihot esculenta) The cassava plant is a staple food crop for over 500 million people in Asia, Africa, South America, and the Caribbean [11]. It is also the primary source from which plant starches, such as tapioca and farina, are derived. From 1985 to 1996, the world production of cassava increased from 134 to 164 million tons per year, an increase of about 2.0% per year [12]. With projected increases in world population, cassava production and usage will likely continue to increase. The HCN content of cassava can reach levels toxic to humans (Table 3.2). However, the HCN content of the bitter species of cassava protects the plant from pests and disease, making it more desirable for cultivation [13]. Consumption of bitter cassava by humans and livestock requires thorough processing to produce safe food products. Safe processing involves the following steps [13]: • Peeling of the root tubers; for animal food the peeling step is omitted • Milling of the peeled tubers (which initiates the HCN release) • Heating and drying to remove the HCN by volatilization In the traditional processing for human consumption, the milled tubers may be washed with water, or sun-dried to remove HCN, or fermented [13]. Traditional processing methods allow the volatilization and release of HCN to the atmosphere, a process that takes days to complete. It is conceivable that as processing moves towards more rapid methods (e.g., extraction in water), the extraction water may pose environmental problems. Similarly, the volume/quantity of peelings, if not reused as animal feed after processing, may also pose disposal problems. Human ingestion of improperly processed bitter cassava has severe consequences, and can result in epidemic spastic paraparesis, more commonly referred to as Konzo disease. The clinical features of the disease are characterized by an abrupt onset of a permanent, symmetrical, but nonprogressive

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TABLE 3.3 Examples of Plant Cyanogenic Glycosides, their Sources, and the Products from their Hydrolysis Glycoside

Plant sources

Amygdalin

Vicianin Dhurrin

Members of the Rosaceae, including almond, apple, apricot, cherry, peach, pear, plum, and quince Members of the Rosaceae, including cherry laurel; Eucalyptus cladocalyx; Linaria striata Dc. Sambucus niagra L. (elderberry), Acacia sp. (Australian acacias) Vicia sp. (common vetch) Sorghum sp. (sorghums, Kaffir corns)

Taxiphyllin

Taxus sp.

Linamarin

Lotaustralin Acacipetalin

Phaseolus lunatis L. (lima bean, many varieties); Linum usitatis-simum L. (linen flax); Manihot sp. (cassava or manioc); Trifolium repens L. (white clover); Lotus sp. (trefoils); Dimorphotheca sp. Occurs with linamarin Acacia sp. (South African acacias)

Triglochinin

Triglochin martimum L. (arrow grass)

Prunasin

Sambunigrin

Hydrolysis products Gentiobiose + HCN + benzaldehyde

D-Glucose + HCN + benzaldehyde

D-Glucose + HCN + benzaldehyde Vicianose + HCN + benzaldehyde D-Glucose + HCN + p-hydroxybenzaldehyde D-Glucose + HCN + p-hydroxybenzaldehyde D-Glucose + HCN + acetone

D-Glucose + HCN + 2-butanone D-Glucose + deimethylketone cyanohydrin D-Glucose + HCN + triglochinic acid

Sources: Data from Conn, E.E., Toxicants Occurring Naturally in Foods, National Academy of Science, Washington, DC, 1973, Chapter 14; Howe, R.H., Proceedings of the Conference on Cyanide and the Environment, Tucson, AZ, 1984, p. 331.

paralysis of both legs in previously healthy persons, and in severe cases, damage to arms and cranial nerves also may occur [14]. These outbreaks of Konzo epidemics have occurred in several African nations, including Mozambique, which experienced over 1000 cases in 1981 in Nampula province, 171 cases in 1988 in Namapa and Erati provinces, and large outbreaks in Mogincual provinces. In 1992 and 1993, outbreaks were also observed in the savanna zone of Bandundu region in Zaire, where 78 cases of Konzo were found in a population of 1936 inhabitants [15]. Mammals, including humans, can ingest sublethal quantities of HCN without adverse effects, as the HCN is enzymatically detoxified to thiocyanate and excreted in the urine [16,17]. Konzo symptoms arise with the rapid ingestion of large quantities of HCN in improperly processed cassava; ingestion rate exceeds the rate at which the body can safely process the ingested HCN. 3.1.2.2 Impact of Cyanogenesis in Forage Sorghum, sudan grasses, arrow grass, velvet grass, white clover, and alfalfa are all important forage crops used in livestock agriculture. These plants are all generally cyanogenic in nature, and can potentially be a hazard for grazing livestock. However, different animal species react differently after

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(a)

(b)

FIGURE 3.2 Biosynthetic pathway for cyanogenic glycoside. (a) Generalized biosynthesis of cyanogenic glycosides. (Source: Lechtenberg, M. and Nahrstedt, A., Cyanogenic Glycosides, John Wiley & Sons, 1999. Reproduced with permission.) (b) Biosynthetic pathway for dhurrin from L-tyrosine. (Source: Halkier, B.A. et al. Cyanide Compounds in Biology, John Wiley & Sons, 1988. Reproduced with permission.)

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ingesting plant material containing cyanogenic glycosides (whole intact plant material) or HCN (plant material which has been stressed, chopped, etc.). These differences are mainly due to differences in anatomical structure and different detoxifying abilities in different animals. Monogastric animals (e.g., horses and swine) have strong hydrochloric acid in their stomachs; this hydrochloric acid tends to react with HCN, liberated by the degradation of cyanogenic glycosides, to yield formic acid and ammonium chloride, substances that are essentially nontoxic. On the other hand, multistomached ruminant animals (e.g., cattle, sheep, and goats) have stomachs operating in the neutral pH range, potentially ideal for absorption of HCN released from the breakdown of cyanoglycosides. It must be recognized that toxic episodes will depend on the rates of ingestion, the quantity of cyanoglycosides or HCN ingested, and the capacity of the animal to oxidize the HCN to thiocyanate. If the rates of ingestion and cyanide concentration are greater than the detoxification rate, then a fatal event will occur [18]. It is estimated that a 450-kg cow should be able to detoxify HCN at a rate of 0.5 g h−1 , and rapid ingestion of 1 g would be fatal within 15 to 20 min [19]. One gram of HCN can be contained in about 2 kg of forage, with a cyanide concentration of 500 mg kg−1 . The cyanide content of forage depends on the following factors [19]: • Species and variety of forage crops: Sorghum generally has much higher cyanogenic content than sudan grass, and the sorghum–sudan grass hybrids contain higher cyanogenic material than sudan grass. • Stage of crop growth: Young growing tips, or shoots or leaves, tend to have the highest cyanogenic contents; mature plant material (e.g., leaves, stalks, etc.) tends to have the lowest content. Sudan grass should not be grazed or green chopped until it reaches a height of at least 45 to 50 cm, and sorghum–sudan grass hybrids should not be grazed or green chopped until they have reached at least 60 to 75 cm in height. Sorghum is generally unsafe for pasture or green chopping until full maturity. Greater details on forage management can be obtained in bulletins from any of the U.S. state universities that provide agricultural extension service.

3.1.3 ETHYLENE PRODUCTION IN PLANTS In addition to the production of cyanogenic glycosides, plants that synthesize ethylene via the enzyme 1-aminocyclopropane-1-carboxylic acid (ACC) oxidase also produce HCN as a by-product [20]. Ethylene is produced by almost all parts of higher plants, and is essential for seed germination, root and shoot growth, flower development, senescence and abscission of flowers and leaves, and ripening of fruit. Ethylene is also produced during periods of stress, including drought, flooding, chilling, exposure to ozone, and mechanical wounding. Figure 3.3 depicts the methionine cycle and ethylene biosynthesis. As shown in this figure, ethylene in higher plants is synthesized from S-adenosyl-L-methionine (SAM) via the intermediate ACC. The last step in the pathway, involving the conversion of ACC to ethylene, requires oxygen, and is catalyzed by the enzyme ACC oxidase. HCN is produced as a final by-product of the ACC oxidase reaction in a stoichiometric 1:1 ratio with ethylene.

3.2 CYANIDE IN MICROORGANISMS Cyanide is produced by, and contained in, various microorganisms. Cyanogenic microorganisms have been identified in the general classes of algae, bacteria, and fungi.

3.2.1 CYANOGENIC ALGAE In a review by Vennesland et al. [21], three species of algae are identified as having the capability to produce HCN. These are Chorella vulgaris and cyanobacteria (blue-green algae) Anacystis nidulans,

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Methi (Yang) cycleonine SAM synthetase

MTA nucleosidase

ACC synthetase

O H N

H2C

C

CoASH CH2

Malonyl-CoA

C

C COO–

H2C

+ NH3

H2C

COO– ACC N-malonyltransferase

N-malonyl-ACC

H2C

COO–

1-Aminocyclopropane1-carboxylic acid(ACC)

ACC oxidase H

H C

C

H ½ O2

CO2 + HCN + H2O

H Ethylene

FIGURE 3.3 Production of hydrogen cyanide during ethylene synthesis. (Source: Adapted from Buchanan, B.B., Biochemistry and Molecular Biology of Plants, American Society of Plant Biologists, Rockville, MD, 2002.)

and Nostroc muscorum. These organisms produce HCN from an aromatic amino acid preursor (e.g., histidine, tryptophane, phenylalanine, and tyrosine) via an amino acid oxidase–peroxidase enzyme system in an oxic and illuminated environment. Histidine produces the greatest amount of HCN. The reaction occurs in a sequence of stages, as shown in Figure 3.4. In the first stage, oxidation of histidine is catalyzed by amino acid oxidase to an imino acid; oxygen is reduced to hydrogen peroxide. In the second stage, imino acid oxidation occurs, and is catalyzed by peroxidase to yield HCN and imidazole aldehyde [21]. It must be noted that in the absence of peroxidase, the imino acid is rapidly hydrolyzed, nonenzymatically, to keto acid and ammonia [21].

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Histidine

FIGURE 3.4 Production of hydrogen cyanide from histidine in amino acid oxidase–peroxidase system. (Source: Reprinted from Cyanide in Biology, Vennesland, B. et al., 349, 1981. With permission from Elsevier.)

2H+ H2N - CH2 - COOH

HN

2H+ CH - COOH

N

≡ C - COOH

HCN + CO2

FIGURE 3.5 Production of hydrogen cyanide from glycine. (Source: Adapted from Knowles, C.J., Bacteriol. Rev., 40, 652, 1976.)

3.2.2 CYANOGENIC BACTERIA Cyanogenesis in bacteria is limited to Chromobacterium violaceum and certain pseudomonad species (i.e., P. chlorophis, P. aureofaciens, P. aeruginosa, and P. fluorescens) [22–32]. These organisms require glycine for the production of HCN, which occurs only in the transition stage of growth, from log phase to stationary phase, under aerobic conditions [22,23,25,26,28,30]. A two-step oxidative reaction model involving two flavoproteins has been proposed for the production of HCN from glycine. This reaction sequence is shown in Figure 3.5, which entails the oxidation of glycine to iminoacetic acid and subsequent oxidation to HCN and carbon dioxide. The HCN in this figure originates from the methylene group, and carbon dioxide from the carboxyl group of glycine [26].

3.2.3 CYANOGENIC FUNGI Formation of HCN was first observed in the fungus Marasmius oreacles in 1871 [25]. Numerous species of fungi have now been identified as cyanogenic. These organisms include species of the genera Actinomycetes, Basidiomycetes, Clitocybe, Marasmius, Pholiota, Polyporus, and Tricholoma [25]. The production of cyanide by fungi varies, depending on the growth stage and species

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TABLE 3.4 Listing of Plant Pathogenic Fungi Organism Basidiomycetes species Marasmium oreacles Stemphylium loti

Disease Snow mold disease Fairy ring disease Copperspot disease

Plants affected Alfalfa and other forage plants Grasses (lawns and pasture) Birdsfoot trefoil

Sources: Data from Castric, P.A., Castric, K.F., and Meganathan, R., Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 236; and from Knowles, C.J., Bacteriol. Rev., 40, 652, 1976.

of the organism. In Pholiota aurea, HCN was formed in young, fresh, fruiting bodies; old fruiting bodies and old mycelia produce HCN only when damaged or stressed [33]. In snow mold basidiomycetes grown in liquid culture, HCN was produced during active growth [34]; in a B-type isolate of the same snow mold basidiomycetes, HCN was produced only during the autolytic growth phase [35]. Fungi are also capable of producing cyanogenic compounds that in turn produce free cyanide. The B-type isolate of the snow mold basidiomycetes produced an unstable cyanogenic compound in shake cultures, with increased rate of production observed during active growth. The cyanogenic compound released free cyanide by a nonenzymatic reaction [35], and was identified predominantly as a glyoxylic acid, cyanohydrin [35,36]. These snow mold fungi also produced a β-glucosidase enzyme [36], which could be responsible for HCN release from plant cyanogenic glycosides. (e.g., linamarin, lotasistrain, and amygaldin). Many cyanogenic fungi are plant pathogens. Table 3.4 presents a listing of these pathogenic fungi, the disease caused by each of them, and the plants affected by these pathogens. It is generally believed that the pathogenic effects are partly because of the ability of the organism to release HCN and partly because of the release of HCN from the host plant’s cyanogenic glycosides, when acted upon by the fungal β-glucosidase enzyme. This may be especially true in the case of the snow mold basidiomycetes [25].

3.2.4 ROLE OF CYANOGENIC COMPOUNDS IN MICROORGANISMS The role of cyanogenic compounds in microorganisms is unclear, except in the case of fungi. It is possible to conclude that in algae and bacteria, the cyanogenesis is simply an end result of the metabolic process under opportunistic circumstances, because of the strict environmental and substrate specificity required for HCN production. In addition, the small quantities of HCN produced, measured as nano-moles in reported tests [22,24,27], likely will not be environmentally significant, and may not be an effective antibiotic deterrent in the natural environment. For example, the bacterial species Pseudomonas aeroginosa has the capacity for cyanogenesis, but also has the capacity to degrade cyanide compounds [37]. Thus, there may not be a net release of HCN in the natural environment. Fungi, on the other hand, are recognized plant pathogens, and their use of HCN, either self-excreted or enzymatically induced from the host plant, is a mechanism to provide growth nutrients. In the field of plant pathology, it is generally accepted that certain microorganisms provide some benefit to “host” plants. Species of pseudomonas (including species known to be cyanogenic) have been recognized to be involved in the suppression of plant root pathogens [38]. This control on pathogens may involve the bacterial secretion of secondary metabolites, including antibiotics, siderophores, and HCN. Cyanogenic processes in microorganisms thus have a range of effects, but generally do not seem to pose any threats to ecosystems and the natural environment.

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3.3 CYANIDE IN ANIMALS 3.3.1 CYANOGENIC ANIMALS The production of HCN by animals is restricted almost exclusively to one phylum, arthropods; even within this phylum, composed of 11 extant classes of animals, the phenomenon is restricted to certain members of Chilopoda (centipedes), Diplopoda (millipedes), and Insecta (insects) [39]. Based on the literature, the following arthropods have been identified as being cyanogenic: • Chilopods (centipedes): 7 out of the 3,000 species which make up the class of chilopoda [39] • Diplopods (millipedes): 46 out of the 7,500 species which make up the class of diplopoda [39] • Insecta (insects): 68 out of the 750,000 species which make up the class of insecta [40] Specific class, families, and species name listings for these animals can be found in the reviews by Duffey [39] and Davis and Nahrstedt [40].

3.3.2 BIOSYNTHESIS OF CYANIDE There has been limited study of the biosynthesis of cyanide in arthropods [39], and the studies performed indicate that to a large degree, cyanide biosynthesis processes in two species of millipedes mirror the processes known to occur in plants [39], where a biogenetic precursor (e.g., L-phenylalanine) is processed to form a cyanogenic glycoside (e.g., mandelonitrile), as illustrated in Figure 3.6. Figure 3.6 shows the cyanogenic release of HCN and an aldehyde (e.g., benzaldehyde). Biosynthesis of cyanogenic glycosides in insects is slightly different from biosynthesis in plants, in that these animals appear to have the capacity to sequester the cyanogenic compounds from their food source, and also synthesize these compounds from amino acids such as valine and isoleucine, the precursors to linamarin and lotaustralin [40]. Also, some insects have the biosynthetic ability to assimilate ingested cyanogenic glycosides into other cyanogenic compounds, as shown in Figure 3.7.

Phenylacetonitrile

FIGURE 3.6 Biosynthetic pathway for hydrogen cyanide and benzaldehyde in the polydesmoid millipede. (Source: Duffey, S.S., Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westerley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 385. With permission.)

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FIGURE 3.7 Transformation of plant cyanoglucoside prunasin to (R)-mandelonitrile by the beetle Paropsis atmoria. (Source: Nahrstedt, A. et al., Cyanide Compounds in Biology, Evered, D. and Harnett, S., Eds., John Wiley & Sons, 1988. Reproduced with permission.)

3.3.3 ROLE OF CYANOGENIC COMPOUNDS IN ARTHROPODS The role of cyanogenesis in arthropods is generally accepted as a defensive or antipredator device [21,39,40]. These defensive chemicals are actively secreted when a predator attacks [21]. However, as in plants, there are apparent exceptions to this defensive function. Certain beetle larvae are voracious predators of both cyanogenic and benzoquinone-producing millipedes, and starlings and toads consistently eat millipedes [39]. These phenomena reflect opportunistic traits of nature. Another example is the opportunistic trait of the larvae of some insects to sequester the cyanogenic materials from plants and in turn, use them for their own defense. Further detail on this behavior is provided in Chapter 12.

3.4 CYANIDE IN FOREST FIRES Annually, there are over 42,000 forest fires (wildfires) in the United States [41]. In the summer of 2000, over 7 million acres of land was burned in the western United States. Similar events occur worldwide [42]. These fires occur as a result of natural events (i.e., electrical storms) and man’s activities (e.g., slash and burn, and accidents). Cyanide is produced in forest fires, and is released to the atmosphere [43]. Emissions of cyanide may be exacerbated in some forest fires, where cyanideamended (as an anticorrosion agent) fire retardant is used to assist in the fighting of the fire [44,45]. Usage of these cyanide-treated fire retardants can result in adverse environmental consequences. A forest fire is an uncontrolled thermal event where organic matter is thermally consumed, at times in the presence of abundant oxygen, and at times with limited oxygen. The organic matter naturally contains some nitrogen along with abundant carbon, and the circulating air contains abundant nitrogen. In those regions of a fire where oxygen is limited, there will be opportunity for HCN to form. In fact, this type of thermal process, where organic matter is fired in a furnace, was one of the early processes for producing cyanide salts [46]. Laboratory simulation of the smoldering remains of a fire, which can persist for weeks after the passage of the flame front and underground fires, has demonstrated the production of HCN methane, ethane, ethylene, acetone, acetonitrile, acetylene, propene, formaldehyde, methanol, acetic acid, formic acid, glycolalehyde, phenol, furan, and ammonia [47,48]. Forest fires release significant quantities of cyanide to the atmosphere. Estimates of the amount of HCN and methyl cyanide released to the atmosphere from biomass burning are in the range of

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(1.4 to 2.9) × 1012 g (as N) per year and about 0.5 × 1012 g (as N) per year, respectively [43,49]. It is believed that the oceans of the world provide the primary receiving sinks for this cyanide and in the absence of any reports of adverse effects, one might assume that these ocean sinks [43] are capable of handling this quantity of cyanide at the rates currently produced. At this time it is difficult to estimate the quantity of cyanide released from fire retardant materials used in fighting forest fires. However, given the large number of outbreaks each year, these releases may be significant.

3.5 SUMMARY AND CONCLUSIONS • HCN and cyanogenic compounds are constituents of the earth’s biosphere. Cyanide contains carbon and nitrogen, key components in the evolution of life on this planet. • Cyanogenesis occurs in plants, microorganisms, and animals. It is believed that all plants produce low levels of cyanogenic compounds, and some 2650 species produce significant levels of free cyanide. Many cyanide-producting plant species bear social and economic importance in our daily life, in areas such as food, forage, and horticulture. • In the animal world, arthropods (centipedes, millipedes, moths, and butterflies) use cyanogenesis as a defense device against predators (e.g., venom in the centipedes and millipedes). • The role of cyanogenesis in algae and bacteria in the natural environment is varied and uncertain. Fungi are plant pathogens, and their use of HCN, either self-excreted or enzymatically induced from the host plant, is a mechanism to provide growth nutrients. • Plant pathogenic fungi use HCN, either self-produced or induced from the host plant, to infest pasture and lawn grasses and other forage plants, as a means of securing growth nutrients. Pathogenic fungi cause plant diseases such as snow mold, fairy ring, and copper spot diseases. • The level of naturally occurring cyanide from the plant, microbial, and animal kingdoms does not appear to pose a major environmental threat to ecosystems. However, with ever increasing demands for food, and increased cultivation of cyanide-producing plants, there could be environmental implications in the processing of food plants like cassava. • Forest fires release significant quantities of cyanide to the atmosphere. The oceans of the world appear to provide the primary receiving sinks for this cyanide.

REFERENCES 1. Oparin, A.I., The Origin of Life, Dover Publications, Mineola, NY, 1953. 2. Allen, J. and Strobel, G.A., The assimilation of H14 CN by a variety of fungi, Can. J. Microbiol., 12, 414, 1966. 3. Thatcher, R.C. and Weaver, T.L., Carbon–nitrogen cycling though microbial formamide metabolism, Science, 192, 1234, 1976. 4. Lechtenberg, M. and Nahrstedt, A., Naturally occurring glycosides, in Cyanogenic Glycosides, Ikan, R., Ed., John Wiley & Sons, Chichester, U.K., 1999, Chapter 5. 5. Jones, D.A., Cyanogenesis in animal–plant interactions, in Cyanide Compounds in Biology, Evered, D. and Garnety, S.F., Eds., John Wiley & Sons, 1988, p. 151. 6. Bolhuis, G.G., The toxicity of the cassava root, Netherlands J. Agr. Sci., 2, 176, 1954. 7. Seigler, D.S., Cyanogenic glycosides and lipids: structural types and distribution, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 133. 8. Georgiadis, N. and McNaughton, S., Interactions between grazers and a cyanogenic grass, Cynodon plectostachyus, Oikos, 51, 343, 1988.

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9. Lieberei, R., Relationship of cyanogenic capacity (HCN-c) of the rubber tree (Hevea brailiensis) to the susceptibility to Microcyclus ulei, the agent causing South American leaf blight, J. Phytopath. (Berl), 122, 54, 1988. 10. Jones, D.A., Selective eating of the acyanogenic form of the plant Lotus corniculatus by various animals, Nature, 193, 1109, 1962. 11. Padmaja, G., The culprit in cassava toxicity: cyanogens or low protein? in Consultative Group on International Agricultural Research News Letter, Volume 3, www.worldbank.org/html/cgiar/newsletter/Oct96/ 6cgnews.html, 1996. 12. Gottret, M.V., Cassava, Section 2.1.B, Annual Report 97, Centro International de Agricultura Tropical (CIAT), Cali, Colombia, www.ciat.org/impact/iannual97/lanual97c.htm, 1997. 13. Cereda, M., Processing of cassaca roots in Brazil, Proceedings of the International Workshop on Cassava Safety, ISHS Acta Horticultura, 375, 21, 1994. 14. Howlett, W.P., Konzo: a new human disease entity, Proceedings of the International Workshop on Cassava Safety, ISHS Acta Horticulturae, 375, 32, 1994. 15. Cliff, J.L., Cassava safety in times of war and drought in Mozambique, Proceedings of the International Workshop on Cassava Safety, ISHS Acta Horticulture, 375, 37, 1994. 16. Jones, D.A., Why are so many food plants cyanogenic? Phytochemistry, 47, 155, 1998. 17. Westley, J., Mammalian cyanide detoxification with sulphane sulfur, in Cyanide Compunds in Biology, Evered, D. and Garnety, S.F., Eds., John Wiley & Sons, Chichester, U.K., 1988, p. 201. 18. USAT, Cherry tree leaves killed foals, scientists conclude, USA Today, July 13, 2001. 19. Vough, L.R. and Cassel, E.K., Prussic acid poisoning of livestock, University of Maryland, College of Agriculture, Maryland Cooperative Extension Bulletin, www.inform.um.edu/prussic_acid_poisoning_ of_ livestock.htm, 1987. 20. Yip, W.K. and Yang, S.F., Ethylene biosynthesis in relation to cyanide metabolism, Botanical Bull. Acad. Sin. Taipei, 39, 1, 1998. 21. Vennesland, B., Pistorius, E.K., and Gewitz, H.S., HCN production by microalgae, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 349. 22. Castric, P.A., Hydrogen cyanide, a secondary metabolite of Pseudomonas aeruginosa, Can. J. Microbiol., 21, 613, 1975. 23. Castric, P.A., Influence of oxygen on Pseudomonas aeroginosa hydrogen cyanide synthase, Curr. Microbiol., 29, 19, 1994. 24. Castric, P.A., Castric, K.F., and Meganathan, R., Factors influencing the termination of cynogenesis in Pseudomonas aeruginosa, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 236. 25. Knowles, C.J., Microorganisms and cyanide, Bacteriol. Rev., 40, 652, 1976. 26. Michaels, R. and Corpe, W.A., Cyanide formation by Chromobacterium violaceum, J. Bacteriol., 89, 106, 1965. 27. Nazly, N., Collins, P.A., and Knowles, C.J., Cyanide production by harvested Chromobacterium violaceum, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 289. 28. Niven, D.F., Collins, P.A., and Knowles, C.J., The respiratory system of Chromobacterium violaceum grown under conditions of high and low cyanide evolution, J. Gen. Microbiol., 90, 271, 1975. 29. Wissing, F., Cyanide formation from oxidation of glycine by Pseudomonas species, J. Bacteriol., 117, 1289, 1974. 30. Wissing, F., Growth curves and pH optima for cyanide producing bacteria, Physiol. Plant., 21, 589, 1968. 31. Wissing, F. and Andersen, K.S., The enzymology of cyanide protection from glycine by Pseudomonas species. Solubilization of the enzyme, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 275. 32. Castric, P.A., personal communication, 2003. 33. Bach, E., The agaric Pholiota aurea: physiology and ecology, Dan. Bot. Ark., 16, 1, 1956. 34. Lebeau, J.B. and Dickson, J.G., Preliminary report on production of hydrogen cyanide by a snow mold pathogen, Phytopathology, 43, 581, 1953.

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35. Ward, E.W.B. and Lebeau, J.B., Autocatalylic production of hydrogen cyanide by a certain snow mold fungi, Can. J. Bot., 40, 85, 1962. 36. Stevens, D.L. and Strobel, G.A., Origin of cyanide in culture of psychrophilic basidiomycetes, J. Bacteriol., 95, 1094, 1968. 37. Dhillon, J.K. and Shivaraman, N., Biodegradation of cyanide compounds by Pseudomonas species, Can. J. Microbiol., 45, 201, 1999. 38. O’Sullivan, D.B. and O’Gara, F., Traits of fluorescent Pseudomonas species involved in the suppressions of plant root pathogens., Microbiol. Rev., 56, 662, 1992. 39. Duffey, S.S., Cyanide and arthropods, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westerley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 385. 40. Davis, R.H. and Nahrstedt, A., Cyanogenesis in insects (Chapter 15), in Comprehensive Insect Physiology Biochemistry and Pharmacology, Vol. II, Kerkut, G.A. and Gilbert, L.I., Eds., Pergamon Press, Oxford, 1985, p. 635. 41. Johnson, R., Wildfires, www.geo-outdoors.info/wildfires.htm, accessed: August 18, 2004. 42. Levine, J.S., Global Biomass Burning: Atmospheric, Climatic, and Biospheric Implications, MIT Press, Cambridge, MA, 1991. 43. Li, Q., Jacob, D.J., Bey, I., Yantosca, R.M., Zhao, Y., Kondo, Y., and Notholt, J., Atmospheric hydrogen cyanide (HCN): biomass burning source, ocean sink? Geophys. Res. Lett., 27, 357, 2000. 44. Milstein, M., Forest service tells fire retardant maker to remove cyanide, The Oregonian, September 20, 2000. 45. Marshall, P., Red rain — effective? Yes. Toxic? Probably, Forest Magazine, http://www.fseee.org/ forestmag/0303redrain.shtml, accessed: April 23, 2005. 46. Robine, R. and Lenglen, M., The Cyanide Industry, John Wiley & Sons, New York, 1906. 47. Bertschi, I., Yokelsom, R.J., Ward, D.E., Babbitt, R.E., Susott, R.A., Goode, J.G., and Hao, W.M., Trace gas and particle emission from forest fires in large diameter and below ground biomass fuels, J. Geophys. Res., 108, 8472, 2003. 48. Holzinger, R., Warneke, C., Jordan, A., Hansel, A., Lindinger, W., Sharffe, D.H., Schade, G., and Crutzen, P.J., Biomass burning as a source of formaldehyde, acetaldehyde, methanol, acetone, acrylonitrile, and hydrogen cyanide, Geophys. Res. Lett., 26, 1161, 1999. 49. Mauresberger, A., Methyl cyanide (CH3 CN) and hydrogen cyanide (HCN): tracers for biomass burning, Max-Planck Institute, Germany, http://mpi-hd.mpg.de/mauersberger/arnold/biomass.htm, accessed: April 23, 2005. 50. Shibamoto, T. and Bjeldanes, L.F., Introduction to Food Toxicology, Academic Press, San Diego, CA, 1993. 51. Tewe, O.O., Detoxification of cassava products and effects of residual toxins on consuming animals, in Proceedings of the FAO Expert Consultation on Roots, Tubers, Plantains and Bananas in Animal Feeding, Machin, D. and Nyvold, S., Centro International de Agricultura Tropical, Cali, Colombia, http://www.fao.org/docrep/003/T0554E/T0554E06.html, January 1991. 52. Nartey, F., Cyanogenesis in tropical foods, in Cyanide in Biology, Vennesland, B., Conn, E.E., Knowles, C.J., Westley, J., and Wissing, F., Eds., Academic Press, London, 1984, p. 115. 53. Ravindran, V., Preparation of Cassava leaf products and their use as animal feeds, in Proceedings of the FAO Expert Consultation on Roots, Tubers, Plantains and Bananas in Animal Feeding, Centro International de Agricultura Tropical (CIAT), http://www.fao.org/DOCREP/003/T0554E/T0554E00.htm, January 1991. 54. Wheeler, J.L., Implications for domestic animals of cyanogenesis in sorghum forage and hay, in Proceedings of the International Workshop on Cassava Safety, ISHS Acta Horticulturae, 375, 25, 1994. 55. Goodger, J., Characterization of cyanogenesis in Australian eucalyptus, in Proceedings of the Plant Biology, http://www.rycomusa.com/aspp2001/public/P39/0478.html, 2001. 56. Conn, E.E., Cyanogenic glycosides, in Toxicants Occurring Naturally in Foods, National Academy of Science, Washington, DC, 1973, Chapter 14. 57. Howe, R.H., The presence of cyanides in nature, in Proceedings of the Conference on Cyanide and the Environment, Tucson, AZ, 1984, p. 331.

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and the Use 4 Manufacture of Cyanide George M. Wong-Chong, David V. Nakles, and Richard G. Luthy CONTENTS 4.1

Production of Cyanide Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.1 Hydrogen Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.2 Production of Sodium Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.2.1 Global and U.S. Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.2.2 Production Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.3 Production of Ferrocyanides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.4 Production of Acrylonitrile . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.4.1 Global and U.S. Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.4.2 Production Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Incidental Industrial Production of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1 Coking and Gasification of Coal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.2 Blast Furnace Operations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.3 Aluminum Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.4 Municipal Waste and Sludge Incineration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

42 42 42 42 44 48 49 49 50 51 52 52 52 53 53 54

Cyanide, a natural compound found in plants and animals, is believed to be a key component in the origin of life (see Chapter 1) and plays a pivotal role in today’s commerce. It is a basic component in the manufacture of a number of products including synthetic fibers and plastic, gold, agricultural herbicides, fumigants and insecticides, dyes and pigments, animal feed supplements, chelating agents for water treatment, and specialty chemicals and pharmaceuticals [1,2]. Table 4.1 presents a breakdown of the overall industrial use of hydrogen cyanide, including as a feedstock chemical for production of other cyanide compounds, as of 1991. Table 4.2 presents a list of some industries that use cyanide compounds in the manufacturing process, along with the cyanide compounds employed. The cyanide industry traces its history to about 1710 with the discovery of Prussian Blue (or ferric ferrocyanide), an iron cyanide compound, which at that time was used almost exclusively in dyeing [3,4]. However, it was not until about 1885 that substantial commercialization of cyanide, specifically potassium cyanide, occurred with the development of the McArthur-Forest process, known today as the cyanidation process, for the extraction of gold from low-grade ores [3]. This discovery represents a major sustaining factor in today’s cyanide commerce, with about 20%, or an estimated 0.6 million tons, of the worldwide production of cyanide used in mining [5,6]. This chapter discusses the manufacture of cyanide compounds, especially hydrogen cyanide, sodium cyanide, ferrocyanide, and acrylonitrile, as well as the uses of these compounds and their 41

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TABLE 4.1 Use of Hydrogen Cyanide in Manufacturing in the United States (1991 Estimate) Product

HCN used (%)

Adiponitrile for nylon Acetone cyanohydrin for plastics Sodium cyanide Cyanuric chloride for pesticides and agricultural products Chelating agents (e.g., EDTA) Methionine, animal feed Misc.: specialty chemicals and pharmaceuticals

41 28 13 9 4 2 3

Source: Data from Pesce, L.D., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993.

production rates. The chapter also discusses those industries where cyanide production is an incidental occurrence, such as in coking and gasification of coal, metal ore reduction in blast furnaces, the reduction of alumina, and municipal waste and sludge incineration.

4.1 PRODUCTION OF CYANIDE COMPOUNDS 4.1.1 HYDROGEN CYANIDE In 2001, the worldwide production of hydrogen cyanide was approximately 2.6 million tons [6]. The U.S. production in the period 1983 through 2001 was 0.33 to 0.75 million tons per year, as shown in Table 4.3. There are four commercial processes for the production of hydrogen cyanide. Two of these are synthesis processes involving the reaction of ammonia, methane (natural gas), and air over a platinum catalyst: (1) the Andrussow process and (2) the Blausaure–Methan–Ammoniak (BMA) process. A third process, the Shawinigan process, uses a carbon fluid bed in an electrical fluohmic furnace to react ammonia and propane. The fourth process is the acrylonitrile production process where hydrogen cyanide is produced as a by-product and which accounts for about 30% of worldwide supply [2]. Table 4.4 presents summary information about the synthesis processes for hydrogen cyanide. The Andrussow process, which is by far the dominant manufacturing process, produces hydrogen cyanide via the following reaction [2]: CH4 + NH3 + 1.5O2 → HCN + 3H2 O

(4.1)

Figure 4.1 presents a schematic flow diagram of the Andrussow process. This diagram shows the recovery/recycle of ammonia and waste heat-design features that improve the efficiency and economy of the process. Details of the process are available in the Kirk–Othmer Encyclopedia of Chemical Technology [2].

4.1.2 PRODUCTION OF SODIUM CYANIDE 4.1.2.1 Global and U.S. Production The McArthur-Forest patent for gold extraction from ore with cyanide was issued in 1887 and the cyanidation process was first used in the Crown Mine in New Zealand and then elsewhere in the

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TABLE 4.2 Use of Cyanide Compounds in Manufacturing Industries Industry Adhesives Cement stabilizer Electroplating

Fire retardant Herbicides Fumigant, poison gas, pesticides, insecticides, parasiticide

Mining

Petroleum Photography

Pharmaceuticals (includes antibiotics, steroids, prescription and nonprescription drugs)

Primary cyanide compounds used in the process Ammonium thiocyanate Calcium cyanide Potassium- or sodium-cyanide (degreasing) Propionitrile (solvent, dielectric fluid) Nickel cyanide Silver cyanide Barium cyanide Zinc cyanide Copper cyanide Hydrogen cyanide Cyanogen chloride (metal cleaner) Mercuric potassium cyanide (mirror manufacturing) Potassium ferrocyanide Ammonium thiocyanate Cyanogen Cyanogen chloride Cyanogen bromide Zinc cyanide Copper cyanide Calcium cyanide Hydrogen cyanide Ammonium thiocyanate (pesticides) Sodium cyanide Malononitrile Cyanogen bromide Barium cyanide Calcium cyanide Ferrocyanide (used as a flotation agent for copper and lead/zinc separation) Malononitrile (lubricating oil additive) Propionitrile (solvent) Ferricyanide bleach Mercuric cyanide Hydrogen cyanide Ferricyanide Ferrocyanide Propionitrile Ammonium thiocyanate (ingredient in antibiotic preparations)

References [14] [15] [14–18]

[19,20] [14,21] [14,15]

[14–17]

[15] [17,22–24]

[14,15,22,24]

(continued)

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TABLE 4.2 Continued Industry Pigments, paints, dyes, ink, personal care products

Road salt

Rocket and missile propellant Synthetic fiber, acrylic fiber, nylon, synthetic rubber

Wine

Primary cyanide compounds used in the process Ferricyanide Ferrocyanide Ferric ferrocyanide (Prussian blue, Fe4 (Fe(CN)6 )3 ) Malononitrile Mercuric cyanide (germicidal soap) Copper cyanide (marine paint) Sodium ferrocyanide Ferric ferrocyanide (Prussian blue, Fe4 (Fe(CN)6 )3 ) Potassium ferrocyanide Cyanogen Ammonium thiocyanate Malononitrile Adiponitrile (intermediate in the manufacture of nylon) Cyanogen bromide Cyanogen chloride Hydrogen cyanide (production of nylon and other synthetic fibers and resins) Ammonium thiocyanate (improve the strength of silks) Potassium ferrocyanide

References [15,25–27]

[17,28–30]

[14,15] [14–16,31]

[32]

Source: Data from MPI, Final Technical Memorandum: Summary of cyanide investiation at SRWTP and preliminary conclusions and recommendations, report by Malcolm Pirnie, Inc., Emeryville, CA to the Sacramento Regional County Sanitation District, Sacramento Regional Wastewater Treatment Plant, Regulatory Compliance Group, Sacramento, CA, 2004.

1890s. This process started the new field of hydrometallurgy. With the advent of this process, world production of potassium cyanide rose from 5,900 tons per year in 1899 to 21,000 tons per year in 1915 [2,3]. Sodium cyanide eventually replaced the potassium salt for economic reasons, and has been the cyanide salt used in hydrometallurgical gold extraction solutions for many years. Production and use of sodium cyanide has been growing. Global annual usage of sodium cyanide in 1989 was about 340,000 tons. In the early 1990s, the total world production of sodium cyanide was estimated to be in excess of 450,000 tons. In 2001, the global production rate was about 600,000 tons per year [2,6]. 4.1.2.2 Production Methods In 1906, Robine and Lenglen [3] cited 79 processes for the production of potassium cyanide: 10 processes involving extraction from ferrocyanide; 13 processes involving extraction from thiocyanate; 28 processes involving synthesis from atmospheric nitrogen; 24 processes involving synthesis from ammonia; and four other processes. In 1891 through 1899, the Beilby process — involving synthesis from ammonia, sodium and potassium carbonate, and powdered charcoal — accounted for about 50% of the total European production of alkali cyanide [2]. In 1900, the Castner

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TABLE 4.3 Production of Hydrogen Cyanide in the United States, 1983–2001 Year

Productiona , 103 tons/yr

2001 2000 1999 1998 1997 1996 1995 1994 1993 1992 1991 1990 1989 1988 1987 1986 1985 1984 1983

750 765 745 725 710 695 675 645 600 570 565 585 565 500 470 430 365 365 330

a

Production estimates for 1983–1988; Source: Data from Pesce, L.D., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. Production estimates for 1989–2001; Source: Data from Myers, E., American Chemistry Council, Washington, DC, personal communication, 2002.

TABLE 4.4 Synthesis Processes for Hydrogen Cyanide Process Andrussow Blausaure–Methan–Ammoniak Shawinigan Acrylonitrile process

Catalysts

Temperature, ◦ C

Feed

Platinum/rubidium Platinum Carbon fluid bed in a fluohmic furnace By-product

1100 1100 1350–1650 400–510

NH3 , air, and CH4 NH3 and CH4 NH3 and C3 H8 NH3 , air, and C3 H6

Source: Data from Pesce, L.D., Kirk-Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993.

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Wastewater

NH3 Recycle

NH3 Fractionator

Waste-Heat Boiler

NH3 / Water

NH3 Feed Air Feed

Reactor

Steam NH3 Absorber Diammonium Phosphate Solution

Natural Gas Feed

NH3 Stripper

Off-Gas Minus NH3

Monoammonium Phosphate Solution

Acid Waste Gases to Boiler or Flare

HCN Absorber

HCN/Water Coolers

Waste Water

HCN Stripper

Steam

HCN/Water

HCN Fractionator

SO2

HCN with SO 2 Inhibitor

FIGURE 4.1 Schematic flow diagram of the Andrussow hydrogen cyanide production process. (Source: Pesce, L.D., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. Reprinted with Permission of John Wiley & Sons, Inc.)

process replaced the Beilby process and dominated production through 1960 for both potassium and sodium cyanide. For the production of sodium cyanide, the Castner process employs elemental sodium and a reaction with ammonia and carbon as follows: 2Na + 2NH3 + 2C → 2NaCN + 3H2

(4.2)

Low yields and elevated costs led to the obsolescence of the Castner process. This process was replaced by the neutralization or wet processes that react hydrogen cyanide from the Andrussow or BMA processes with a sodium hydroxide solution: HCN + NaOH → NaCN + H2 O

(4.3)

Most modern, high tonnage production plants use essentially purified anhydrous liquid hydrogen cyanide to react with sodium hydroxide to produce a product consisting of 99% sodium cyanide. The manufacturing process includes the evaporation of water and crystallization of the sodium cyanide. Control of the process is critical to maximize the average crystal size; to avoid hydrogen

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Dehumidifier Scrubber

Vacuum system

Dust Scrubber

Condenser

Waste Cyclone Separator Air Heater

50% Caustic

Hydrogen cyanide

Crystallizer System

Filter

Briquetter

Mixing Conveyor Steam Screens

Product to packaging and storage

FIGURE 4.2 Production process flow diagram for sodium cyanide. (Source: Pesce, L.D., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. Reprinted with permission of John Wiley & Sons, Inc.)

cyanide polymer formation, which produces an off-white product; and to minimize sodium formate formation, which reduces product purity. Figure 4.2 presents a process flow diagram for a typical sodium cyanide production plant. An occasionally used, alternative process is the direct absorption of crude hydrogen cyanide gas from the manufacturing operation into a sodium hydroxide solution. However, the purity of the sodium cyanide product is lower, that is, approximately 96 to 97% [2]. The primary impurities are sodium carbonate and sodium formate. The formation of larger crystals facilitates the dewatering in the filtration step. In many plants, the moist salt from the filter is passed through a mixing conveyor to destroy the lumps. Often, heated air (450◦ C) is passed through the cake on the filter and through the mixing conveyor. Drying is completed in a hot-air conveyor-dryer. This approach to drying avoids the overheating of the sodium cyanide crystals, thus minimizing the formation of sodium formate in the dried product. A slight excess of sodium hydroxide must be maintained at all stages of processing to maintain an elevated pH, which

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prevents the formation of a black or brown hydrogen cyanide polymer. The product, as shipped, must also contain a slight excess of sodium hydroxide, to ensure that the product yields clear solutions following arrival at its destination. Inherently, the sodium cyanide forms a 50-µm diameter crystal, yielding a dusty solid of low bulk density that must be compacted or fused into larger particles for safer handling. Due to the expense associated with melting the product for casting it in molds, most processes employ mechanical compacting devices that produce either briquettes or granular products. The compaction process occurs using heat and pressure. Most sodium cyanide is sold in dry form to minimize transportation costs although appreciable tonnage is also sold as a 30% aqueous solution [2]. About 90% of today’s sodium cyanide is used in gold extraction [5,6]. Plants for the production of sodium cyanide, using these processes, are operating in the United States, Italy, Japan, the United Kingdom, Australia, Germany, and China.

4.1.3 PRODUCTION OF FERROCYANIDES Ferric ferrocyanide, also known as Prussian Blue (Fe4 [Fe(CN)6 ]3 ), was the first cyanide compound put to commercial use. The compound was discovered by a Berlin color maker in 1704 [3]. This led to a long history of ferrocyanide chemistry, which has resulted in the use of these compounds in a wide variety of industrially significant applications. A treatise on the chemistry of ferrocyanides [7] describes some 22 applications, and these are listed in Table 4.5. In the late 1700s through the early 1900s, ferrocyanide salts were produced by (1) the synthetic fusion of nitrogenous organic residues (e.g., animal blood, hides, hornes, waste/scrap leather, etc.), potash, and iron, and (2) the direct extraction from illuminating-gas and from the by-product

TABLE 4.5 Uses of Ferrocyanides and their Derivatives in Industry Analytical chemistry Anticaking agent Blueprints Case hardening and heat treatment of steel Chemical synthesis: catalysts, reaction intermediates, and reagents Chemotherapy Corrosion inhibitors Desulfurization of coke oven gas Detergents Dying of textiles Electrical equipment treatment: corrosion resistance; arc stabilization and lowering of grounding resistance Electroplating Minerals dressing, beneficiation, and mining Pesticides Petroleum refining: trace metals removal Photography Pigments and dyes Pickling of steel Rubber: peptizing agent, stabilization agent, and accelerator Separation and identification of organic bases Trace metals removal in fermentation Source: Data from ACC, The Chemistry of the Ferrocyanides, American Cyanamid Co., New York, NY, 1953.

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of illuminating-gas clean-up (e.g., spent iron oxide boxes for gas purification) [3]. It was estimated that about 1.8% of the nitrogen in coal reacted to form hydrogen cyanide during coal gasification. In the direct gas extraction processes, the illuminating gas was scrubbed with an alkaline iron salt solution. Robine and Lenglen [3] discussed in detail nine processes for the direct extraction of cyanide from illuminating gas, three processes for extraction from ammoniacal liquor, and 11 processes for recovering ferrocyanide from spent iron oxide. In the synthesis from nitrogenous organic matter, the process chemistry for making potassium ferrocyanide was thought to be: K2 CO3 + Nitrogenous Matter + Energy → KCN + · · ·

(4.4)

6KCN + Fe2+ → K4 Fe(CN)6 + 2K +

(4.5)

In the first reaction, hydrogen cyanide is produced by the thermal breakdown of the organic matter in an oxygen controlled environment (Equation [4.4]). Subsequently, the hydrogen cyanide reacts with potassium to form potassium cyanide. The potassium cyanide, in turn, reacts with the iron to form potassium ferrocyanide as shown in Equation (4.5). Today, ferrocyanide production utilizes the crude sodium cyanide, produced as described in Section 4.1.2, and ferrous sulfate to form sodium ferrocyanide: 6NaCN + FeSO4 + Heat → Na4 Fe(CN)6 + Na2 SO4

(4.6)

The sodium ferrocyanide is recovered by crystallization as the decahydrate salt. The potassium salt is produced by reacting sodium ferrocyanide with calcium hydroxide and potassium chloride and carbonate according to the following reactions: Na4 Fe(CN)6 + 2Ca(OH)2 → Ca2 Fe(CN)6(s) + 4Na(OH)

(4.7)

Ca2 Fe(CN)6 + 2K2 CO3 → K4 Fe(CN)6 + CaCO3(s)

(4.8)

In earlier times, ca. 1900, Prussian Blue was produced in a two stage process. The first stage reacted potassium ferrocyanide and ferrous sulfate to form a grayish-white precipitate of potassium ferric–ferrocyanide. In the second stage, the potassium ferric–ferrocyanide is oxidized to the tetrairon(III) tris(hexakiscyanoferrate), Fe4 [Fe(CN)6 ]3 [3]. Today, the production of Prussian Blue is more direct, where ferrocyanide is reacted with excess iron(III) to produce the intense blue precipitate [2].

4.1.4 PRODUCTION OF ACRYLONITRILE Acrylonitrile [C3 H3 N], also called vinyl cyanide, is among the top 50 chemicals produced in the United States as a result of the tremendous growth in its use as a starting material for a wide range of chemical and polymer products. Acrylic fibers remain the largest use of acrylonitrile. Other significant uses are resins and nitrile elastomers and as an intermediate in the production of adiponitrile and acrylamide. 4.1.4.1 Global and U.S. Production Worldwide production of acrylonitrile was approximately 3.2 million tons in 1988 [8]. As shown in Table 4.6, more than one-half of that production was located in Western Europe and the United States. In the United States, BP Chemicals dominated production, supplying more than one-third of domestic production. Nearly one-half of the United States production was exported in 1988, with most going to Japan and the Far East [8]. This export market grew steadily from the mid-1970s

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TABLE 4.6 Worldwide Acrylonitrile Production, 1988 Production, 103 tons

Region Western Europe United States Japan Far East Mexico

1200 1170 600 200 60

Total

3230

Source: From Brazdil, F., Kirk-Othmer Encylopedia of Chemical Technology, Vol. 1, John Wiley & Sons, New York, 1993.

TABLE 4.7 Worldwide Acrylonitrile Demand, 103 Tons per Year Region Western Europe Japan United States Far East Mexico/South America Total

1976

1980

1985

1988

880 570 590 200 81

880 510 660 270 130

1140 635 640 385 200

1200 680 660 560 250

2321

2450

3000

3350

Source: Data from Brazdil, F., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 1, John Wiley & Sons, New York, 1993.

to 1988. During this period, it increased from 10% in the mid-1970s to 53% and 43% in 1987 and 1988, respectively. The large exports to the Far East were the result of higher raw material costs (i.e., propylene costs) relative to the United States. A more detailed breakdown of the world demand for acrylonitrile for the period between 1976 and 1988 is provided in Table 4.7. Growth in demand during this period averaged about 3.6% per year between 1984 and 1988. Projections beyond 1988 were 3% per year through 1993.

4.1.4.2 Production Methods Prior to 1960, processes based on either ethylene oxide and hydrogen cyanide or acetylene and hydrogen cyanide were used to produce acrylonitrile. Growth in the demand for acrylic fibers around 1950 spurred improvements in process technology and resulted in the discovery of a heterogeneous vapor-phase catalytic process. This process, which produced acrylonitrile using selective oxidation of propylene and ammonia, is commonly referred to as the propylene ammoxidation process. This process was introduced in 1960 and eventually displaced all other acrlyonitrile manufacturing processes. As of 1988, over 90% of the approximately 3.2 million metric tons of acrylonitrile produced worldwide each year was manufactured using the propylene ammoxidation process [8].

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Fluid-Bed Catalytic Reactor 400° - 510° C 49-196 kPa

Absorber

51

Acrylonitrile recovery column

Acetonitrile recovery column

Lights column

Product column

Crude acrylonitrile Crude acetonitrile

Off-Gas

H.P. Steam

Crude HCN

Product acrylonitrile

H2O

BFW

Air Ammonia START

H2O

Proplene

Heavy impurities

FIGURE 4.3 Process flow diagram for ammoxidation process. (Source: Brazdil, F., Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 1, John Wiley & Sons, New York, 1993. Reprinted with permission of John Wiley & Sons, Inc.)

The primary chemical reaction of the propylene ammoxidation process is as follows: C3 H6 + NH3 + 1.5O2

CATALYST

−→

C3 H3 N + 3H2 O

(4.9)

A process diagram of the commercial process is shown in Figure 4.3. This process uses a fluidized bed reactor, in which propylene, ammonia, and air contact a solid catalyst at 400 to 510◦ C and 49 to 196 kPa gauge. It is a single pass process that achieves about 98% conversion of the propylene, and uses about 1.1 kg of propylene per kilogram of acrylonitrile produced. As shown in Figure 4.3, hydrogen cyanide is a by-product of the acrylonitrile production process. This hydrogen cyanide can be processed as a salable product or used in the manufacture of methyl methacrylate and acetonitrile, common industrial solvents [8].

4.2 INCIDENTAL INDUSTRIAL PRODUCTION OF CYANIDE Many industrial operations that employ thermal processing of carbonaceous materials produce small quantities of cyanide. Included among these operations are • • • •

Coking and gasification of coal Blast furnace processing for iron and nonferrous metal oxide reduction Alumina reduction Municipal waste and sludge incineration

Brief discussions of these operations follow. More details and discussions of these operations are presented in Chapter 26.

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4.2.1 COKING AND GASIFICATION OF COAL The coking operation involves distillation of coal by indirectly heating the coal in the absence of air to temperatures in the range of 900 to 1100◦ C to vaporize all volatile constituents in the coal [9]. These volatile constituents include a range of hydrocarbons (e.g., benzene, toluene, methane, naphthalene, phenols, xylenes, and polynuclear aromatic hydrocarbons), nitrogenous compounds including ammonia, and sulfurous compounds including hydrogen sulfide. The heating period can range from 18 to 36 h [9]. Hydrogen cyanide is formed, albeit in relatively small amounts, due to the high temperature, reducing atmosphere, and the presence of nitrogen and carbon. The quantity of cyanide produced in the coking of coal has been reported to be about 1.5 to 2.0% of the nitrogen content of the coal [3]. A portion of the cyanide remains in the coke oven gas while the remainder leaves the coking system in the waste ammonia liquor wastewater [3]. In the early days of the cyanide industry, ca. 1900, cyanide was recovered from illuminating-gas production, which was essentially a coking process [3,10]. The cyanide was recovered by absorbing the hydrogen cyanide from the gas stream into an alkaline iron salt solution to form an alkali ferrocyanide product. As previously noted, Robine and Lenglen [3] provide detailed descriptions of nine processes for the direct extraction of hydrogen cyanide from illuminating gas and three processes for the extraction of cyanide compounds from the ammoniacal liquors. The coal gasification process is similar to coking, in that the coal is heated to similar temperatures. However, air is introduced into the coal gasifier to combust a portion of the coal, which produces the heat for the steam/coal reactions that take place. The quantity of air that is injected is controlled to maintain the gasification temperature and the quality of product gas that is produced (high or low BTU content). Again, similar to the coking process, some hydrogen cyanide is formed given the conditions of high temperature, abundance of carbon (as low molecular weight hydrocarbons, carbon dioxide or carbon monoxide), and nitrogen from the injected air or coal. Just as in the coking operation, some of this cyanide remains in the product gas and the remaining portion exits the plant in the gas cleaning wastewater.

4.2.2 BLAST FURNACE OPERATIONS In blast furnace operations, the furnace is charged with coke, metal oxide, and limestone flux. The furnace is heated to maintain a temperature profile of about 900◦ C in the upper section of the furnace and greater than 1770◦ C in the bottom section around the tuyeres. Periodic blasts of air and supplemental fuel (oil) are fired into the reaction mixture [9]. This environment of high temperature, available carbon as carbon monoxide and carbon dioxide, and nitrogen from the air blast provide conditions for hydrogen cyanide formation. Flue gas emitted from the top of the blast furnace, about 2.5 to 3.5 tons of gas per ton of iron produced, has a heating value of 80 to 90 BTU/SCF [9] and fuels auxiliary stoves or powers boilers and blowers. The blast furnace gas is laden with dust, as much as 0.05 tons per ton of iron produced, that must be removed to prevent clogging of combustion equipment. About 70% of the dust is removed by bag-houses and dry scrubbing and the remaining 30% by wet scrubbing. The hydrogen cyanide that is present in the gas is removed during the wet scrubbing process and reports to the blast furnace gas scrubber water.

4.2.3 ALUMINUM PRODUCTION Aluminum is manufactured via the electrometallurgical reduction of alumina (Al2 O3 (s)) in the Hall– Heroult process [2,11]. The alumina is placed in a molten cryolite (Na 3 AlF6 ) bath in a carbon-lined cell, or “pot.” An electrical current is passed through the reaction mixture using carbon anodes placed in the molten mixture. The carbon pot liner, which typically is 15 in. thick, serves as the cathode. The molten mass attains a temperature of about 950 to 1,000◦ C [2,11]. The carbon anode is rapidly

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consumed while the carbon pot liner cathode is re-used, usually for several years until it becomes unacceptably contaminated [11]. In this reactor, the conditions of high temperature, available carbon, and available nitrogen provide the opportunity for hydrogen cyanide formation. Air entering at reactor seals is the primary source of nitrogen. The hydrogen cyanide produced becomes absorbed into the carbon of the pot liner and at the end of a pot liner processing life, the cyanide concentration can be as high as 0.9% by weight [11]. Cyanide levels vary within a pot, with highest concentrations observed in the potlining at the side walls. Spent potlining containing cyanide and other contaminants is removed for treatment and disposal. Until the early 1970s, spent potlining was managed as an inert residue and was often used as onsite fill material. Today, spent potlining is a listed hazardous waste (K088) in the United States. Based on data from 1991 to 1997, the annual production of K088 waste is approximately 80,000 to 100,000 tons [12]. Past practices for the disposal of spent potlining have resulted in adverse environmental impacts such as contamination of soil and groundwater, as discussed in more detail in Chapter 27.

4.2.4 MUNICIPAL WASTE AND SLUDGE INCINERATION In areas where land-disposal of municipal wastewater treatment sludge is not practical, sludge incineration is an accepted disposal alternative. In some instances, the economics are improved by combined incineration of municipal refuse and sludge coupled with co-generation of electricity. In the northeastern states of Massachusetts and Rhode Island, municipal wastewater sludge incineration is widely practiced, for example, at the Cranston, Rhode Island and Fitchburg, Massachusetts municipal wastewater treatment plants. These sludges contain about 5 to 6% nitrogen on a dry weight basis and are essentially organic in composition. They are charged to the incinerator with about 70 to 80% water. During the course of sludge incineration, there tends to be pockets of reduced conditions, especially in the early stages of incineration. These localized conditions, coupled with the high temperature and availability of nitrogen and carbon, provide the opportunity for hydrogen cyanide formation. This cyanide leaves the incinerator in the exhaust gases and is transferred to the off-gas scrubbing water. At the Cranston, Rhode Island POTW, an average of 2.08 g cyanide per kilogram of dry sludge incinerated is removed in the scrubber water [13].

4.3 SUMMARY AND CONCLUSIONS • Hydrogen cyanide and other cyanide compounds are used extensively in manufacturing, including the production of synthetic fibers and plastics, agricultural herbicides, fumigants and insecticides, dyes and pigments, animal feed supplements, chelating agents for water treatment, and specialty chemicals and pharmaceuticals. • Sodium cyanide is used extensively for the extraction of gold from ore in hydrometallurgical gold mining. • In 2001, the worldwide production of hydrogen cyanide was approximately 2.6 million tons. • Hydrogen cyanide is manufactured primarily by the Andrussow process, which involves the reaction of methane, ammonia, and oxygen. Hydrogen cyanide from this process is reacted with sodium hydroxide solution to form sodium cyanide, the most common solid form of cyanide that is used in commerce. • Potassium ferrocyanide (K4 [Fe(CN)6 ]), and related solid-phase precipitates, especially ferric ferrocyanide (Fe4 [Fe(CN)6 ]3 ), or Prussian Blue, is produced in large quantity for a variety of specialty uses. • Acrylonitrile, also called vinyl cyanide, is produced in large quantities because of its use as a starting material for a wide range of chemical and polymer products.

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• Various cyanide compounds are produced incidentally during the manufacture of coke, steel, and aluminum, and during the incineration of municipal waste and wastewater sludges.

REFERENCES 1. Homan, E.R., Reactions processes and materials with potential for cyanide exposure, in Clinical and Experimental Toxicology of Cyanides, Ballantyne, B. and Marrs, T.C., Eds., John Wright, Bristol, UK, 1987, p. 1. 2. Pesce, L.D., Cyanides, in Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. 3. Robine, R. and Lenglen, M., The Cyanide Industry, John Wiley & Sons, New York, 1906. 4. Ritter, S.K., Prussian Blue still a hot topic, Chem. & Eng. News, May 2, 2005, p. 32. 5. Logsdon, M.J., Hagelstein, K., and Mudder, T.I., The management of cyanide in gold extraction, International Center for Metals and the Environment, Ottawa, Canada, 1999. 6. Young, C.A., Cyanide: just the facts, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals & Materials Society, Warrendale, PA, 2001, p. 97. 7. ACC, The Chemistry of the Ferrocyanides, American Cyanamid Co., New York, NY, 1953. 8. Brazdil, F., Acrylonitrile, in Kirk-Othmer Encyclopedia of Chemical Technology, Vol. 1, John Wiley & Sons, Inc., New York, 1993, p. 352. 9. USS, The Making, Shaping and Treating of Steel, McGannon, H.E., Ed., United States Steel Corp., Pittsburgh, PA 1971. 10. Hayes, T.D., Linz, D.G., Nakles, D.V., and Leuschner, A.P., Eds., Management of Manufactured Gas Plant Sites, Vol. 1, Amherst Scientific Publishers, Amherst, MA, 1996, Chapter 2, p. 5. 11. USEPA, Best demonstrated available technology (BDAT) background document for spent aluminum potliners — K088, U.S. Environmental Protection Agency, Office of Solid Waste, http:// www.epa.gov/epaoswer/hazwaste/ldr/k088/k088back.pdf, 2000. 12. USEPA, Land disposal restrictions — background document to establish the effective date for amended treatment standards for spent aluminum potliners (proposed rule), U.S. Environmental Protection Agency, Washington, D.C., http://www.epa.gov/epaoswer/hazwaste/ldr/k088/landdisp.pdf, 2000. 13. Bratina, C., Cranston, Rhode Island wastewater treatment plant, personal communication, 2004. 14. ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997. 15. USNLM, Toxicology and Environmental Health Information Program. Toxicology Data Network (TOXNET), U.S. National Library of Medicine, http://toxnet.nlm.nih.gov, accessed: February 18, 2005. 16. Boucabeille, C., Bories, A., Olliver, P., and Michel, G., Microbial degradation of metal complexed cyanides and thiocyanate from mining wastewaters, Environ. Pollut., 84, 59, 1994. 17. Kjeldsen, P., Behaviour of cyanides in soil and groundwater: a review, Water, Air, Soil Pollut., 115, 279, 1999. 18. Patterson, J.W., Cyanide, in Industrial Wastewater Treatment Technology, Butterworth, Boston, 1985, p. 115. 19. Hopkins, S.J., Special water quality survey of the Pecos and Gallinas Rivers below the Viveash and Manuelitas fires of 2000, Surveillance and Standards Section, New Mexico Environment Department, 2001, Available at http://www.nmenv.state.nm.us/swqb/viveash_fire_report_02-2001.html, accessed: February 25, 2005. 20. Little, E. and Calfee, R., The effects of UVB radiation on the toxicity of firefighting chemicals, U.S. Department of Agriculture, Forest Service and Wildland Fire Chemical Systems, 2002. 21. Rapean, J.C., Johnson, R.A., and Hanson, T.P., Biodegradation of cyanide: Nitrite interference in the standard test for total cyanide, in Proceedings of the 35th Purdue Industrial Waste Conference, West Lafayette, IN, 1980, p. 430. 22. USEPA, Sustainable industry project phase I report, Chapter 3, The photoimaging industry, U.S. Environmental Protection Agency, Office of Policy Development, Washington, D.C., 2002.

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23. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 24. Owerbach, D., Analysis and sample stability of cyanide in industrial effluents, J. Water Pollut. Control Fed., 52, 11, 1980. 25. USEPA, Seminar publication — National conference on urban runoff management: Enhancing urban watershed at the local, county and state levels, EPA-625/R-95-003, U.S. Environmental Protection Agency, Washington, D.C., 1995. 26. USEPA, Profile of the wood furniture and fixtures industry, EPA-310/R-95-003, U.S. Environmental Protection Agency, Office of Enforcement and Compliance Assurance, Washington, D.C., 1995. 27. USEPA, Pharmaceutical manufacturing category effluent limitations guidelines, Pretreatment standards and new source performance standards, Final Rule, U.S. Environmental Protection Agency, 40 CFR, Parts 136 and 439, 1998. 28. USEPA, Seminar publication — Wellhead protection: a guide for small communities, EPA-625/R-93002, U.S. Environmental Protection Agency, Washington, D.C., 1993. 29. ICF, Construction and demolition waste landfills, Report by ICF, Inc. Fairfax, VA, Contract No. 68-W30008, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, D.C., 1995. 30. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 31. USEPA, Profile of the motor vehicle assembly industry, EPA-310/R-95-009, U.S. Environmental Protection Agency, Washington, D.C., 1995. 32. USEPA, Consumer fact sheet on cyanide, U.S. Environmental Protection Agency, Office of Ground Water and Drinking Water, http://www.epa.gov/ogwdw/dwh/c-ioc/cyanide.html. Accessed: February 25, 2005. 33. MPI, Final Technical Memorandum: Summary of cyanide investigation at SRWTP and preliminary conclusions and recommendations, report by Malcolm Pirnie, Inc., Emeryville, CA to the Sacramento Regional County Sanitation District, Sacramento Regional Wastewater Treatment Plant, Regulatory Compliance Group, Sacramento, CA, 2004. 34. Myers, E., American Chemistry Council, Washington, DC, personal communication, 2002.

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Properties 5 Physical–Chemical and Reactivity of Cyanide in Water and Soil

David A. Dzombak, Rajat S. Ghosh, and Thomas C. Young CONTENTS 5.1

Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.1 Cyanide Ion Bonding . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.2 HCN Formation and Dissociation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.3 HCN Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.4 Free Cyanide Adsorption to Soil and Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.5 Free Cyanide Oxidation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.6 Free Cyanide Hydrolysis. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 Metal Cyanides: Aqueous Species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1 Weak Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1.1 Formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1.2 Dissociation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1.3 Adsorption on Soil and Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1.4 Oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2 Strong Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2.1 Formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2.2 Dissociation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2.3 Adsorption on Soil and Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2.4 Oxidation–Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3 Metal–Cyanides: Solid Phase Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.1 Simple Cyanide Solids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.2 Alkali or Alkaline Earth Metal–Metal Cyanide Complex Solids . . . . . . . . . . . . . . . 5.3.3 Other Metal–Metal Cyanide Complex Solids. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4 Cyanate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5 Thiocyanate. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6 Organocyanides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.7 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

58 58 58 60 61 62 64 65 65 65 67 68 71 73 73 75 76 78 79 80 80 80 82 84 86 88 88

The reactivity, fate, and toxicity of cyanide in water and soil is highly dependent on the chemical speciation of the cyanide. As outlined in Chapter 2, many different soluble and solid forms of cyanide exist. The simplest form of soluble cyanide is the negatively charged cyanide ion, CN− , which is composed of a carbon atom triple bonded to a nitrogen atom (–C≡N). The nature of this triple bond controls the reactivity of the cyanide anion, including complexation with other metal cations, 57

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formation of molecular hydrogen cyanide (HCN), oxidation of cyanide to cyanate, and adsorption onto clays and other soil components. In environmental systems, wastewaters, and wastes, cyanide usually is found in free and complexed forms, as HCN and as metal–cyanide complexes. Because of a reactive electronic arrangement, cyanide anions can readily form metal–cyanide complexes with most metal cations. Most of these complexes exist as soluble species, but many, particularly iron-cyanide complexes, can react further with metal cations to form stable cyanide solids. The soluble and solid phase cyanide species that occur most often in water and soil are outlined in Chapter 2 and examined in more detail here. In this chapter, the specific physical–chemical properties and reactivity characteristics of the different chemical forms of cyanide are presented. Included are examinations of the nature of bonding in and with the cyano group and free cyanide speciation; the properties and reactivities of soluble metal–cyanide complexes; the properties and reactivities of metal–cyanide complex solids; and the properties and reactivities of cyanate, thiocyanate, and organocyanide compounds.

5.1 FREE CYANIDE 5.1.1 CYANIDE ION BONDING Free cyanide consists of the cyanide anion, CN− , and molecular hydrogen cyanide, HCN, both existing as water soluble entities. The cyanide ion acts as a monodentate ligand with the carbon acting as the donor atom, and also as an ambidentate ligand acting as a donor at both ends of the ion [1]. Several structural factors govern the reactivity of free cyanide. The triple bonded structure of a cyanide anion is comprised of a sigma bond, two π bonds, and two empty bonding orbitals [2]. The “s” and the “p” orbitals are filled with maximum number of electrons, while the “d” and “f” orbitals are empty. This configuration allows for a number of bonding arrangements. Since halogens also have filled “s” and “p” orbitals, the behavior of the cyanide anion is similar to that of halogens [3]. The cyanide ion is considered a pseudo-halide in that it can form π -acceptor covalent bonds with transition metals [3]. It may also share electrons at the triple bond with the Group VI elements oxygen and sulfur, forming cyanate, CNO− , or thiocyanate, SCN− [3], or may act as a strong nucleophile in reactions with organic molecules, for example, nucleophilic addition reactions with aldehydes and ketones to form cyanohydrins [4]. The cyanide ion readily forms neutral compounds or anionic complexes with most major metal cations. The partially or wholly filled “d” orbitals of transition series metals can form covalent bonds with the empty anti-bonding orbitals of the cyanide ion. This involves acceptance of electron density into π orbitals of the carbon atom. The cyanide ion is a strong σ donor, which is responsible for the high stability of some of the metal–cyanide complexes [3].

5.1.2 HCN FORMATION AND DISSOCIATION The cyanide anion protonates in water to form hydrocyanic acid, HCN, the most toxic form of cyanide (see Chapters 13 and 14). The pKa for HCN dissociation reaction is 9.24 at 25◦ C [5]. Thus, at pH greater than 9.24, cyanide anion dominates free cyanide speciation, while soluble HCN is the dominant species under acidic to neutral pH conditions (pH < 9.24). The free cyanide dissociation reaction is as follows: HCN = H+ + CN− ,

pK a = 9.24 at 25◦ C,

I=0

(5.1)

Figure 5.1 shows the distribution of HCN and CN− species as a function of pH for a simple aqueous solution at 25◦ C. The temperature dependence of the equilibrium constant governing the species

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1.0 CN–

Ionization fraction ([HCN]/CNT, [CN-]/ CNT)

HCN 0.8

0.6

0.4

0.2

0

5

6

7

8 pH

9

10

11

FIGURE 5.1 Free cyanide species distribution as a function of pH at 25◦ C(pKa = 9.24 for HCN dissociation at T = 25◦ C, I = 0).

distribution of free cyanide can be calculated via the van’t Hoff equation: ln(K2 /K1 ) = (Hr,25C /R)[1/T1 − 1/T2 ]

(5.2)

where Hr,25C is the standard enthalpy change of reaction at 25◦ C(298 K), R is the molar gas constant (8.314 × 10−3 kJ mol−1 K −1 ), T1 is the reference temperature (298 K), and T2 is the temperature of interest in K. The standard enthalpy change for the reaction given in Equation (5.1) is 146 kJ mol−1 , as calculated using the thermodynamic data compiled in Stumm and Morgan [6]. Substitution of this value, and assuming it is approximately constant for the temperature range 5 to 30◦ C, enables calculation of the temperature dependence of the acidity constant in Equation (5.1): KT = exp[1.756 × 104 K(3.356 × 10−3 K −1 − T−1 ) − 21.28]

(5.3)

where KT is the equilibrium constant for HCN dissociation at the temperature T (K) of interest. Combining Equation (5.3) with the mass action equation for the reaction in Equation (5.1), and the mass balance equation for free cyanide (molar concentrations in [ ]), TOTCN = [HCN] + [CN− ]

(5.4)

yields the following expression for the species distribution fractions for HCN and CN− : αHCN = [HCN]/TOTCN = {H+ }/[{H+ } + exp[1.756 × 104 K(3.356 × 10−3 K −1 − T−1 ) − 21.28]] αCN = [CN− ]/TOTCN = 1 − αHCN

(5.5) (5.6)

where {H+ } is the hydrogen ion activity, 10−pH . The species distribution fraction for HCN, αHCN , is presented in Figure 5.2 for temperatures between 5 and 30◦ C (278 and 303 K), and zero ionic

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1 0.9 0.8

T = 5°C T = 25°C

[HCN]/CNT

0.7

T = 15°C

0.6 0.5

T = 10°C

T = 20°C

0.4

T = 30°C

0.3 0.2 0.1 0 6

7

8

9

10

11

pH

FIGURE 5.2 Ionization fraction for HCN as a function of pH and temperature (I = 0).

TABLE 5.1 Literature Values of Henry’s Law Constant (KH,HCN ) for HCN Temp. (◦ C) 25 Not given Not given 25

KH (atm L mol−1 )

−1 KH (mg L−1 a /mg Lw )

0.122 0.073 0.104 to 0.114 0.115

0.005 0.003 0.0043 to 0.0047 0.0047

Reference Bodek et al. [7] Doudoroff [80] Smith and Mudder [2] Avedesian [81]

strength (I). As is evident in Figure 5.2, temperature has a significant effect on free cyanide species distribution. As temperature decreases, dissociation of HCN decreases, extending the species dominance of HCN to higher pH values.

5.1.3 HCN VOLATILIZATION Hydrogen cyanide has a very low boiling point (25.7◦ C) and thus is volatile in water under environmental conditions. The equilibrium air–water partitioning of HCN can be described by Henry’s Law: PHCN = KH,HCN [HCN]

(5.7)

where PHCN is the partial pressure of HCN gas, atm, KH,HCN the Henry’s Law constant, atm L mol−1 , and [HCN] the equilibrium aqueous phase concentration of HCN, mol L−1 . Table 5.1 lists reported values of Henry’s Law constant for HCN. Henry’s Law constants with units relevant to Equation (5.1) are provided, along with dimensionless analogs corresponding to an equilibrium partitioning expression in which both the aqueous and gas phase concentrations are expressed in the same mass concentration units. Note that the Henry’s Law constant is a function of temperature. There are various empirical relationships that express Henry’s Law constant as a function

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of temperature. One such relationship, reported by Bodek et al. [7], is as follows: log KH,HCN = −1272.9/T + 6.238

(5.8)

where KH,HCN is the Henry’s Law constant, mm Hg/M and T the Temperature, K. Equation (5.8) is reported to be valid for HCN concentrations ranging from 0.01 to 0.5 M and temperatures from 20 to 95◦ C.

5.1.4 FREE CYANIDE ADSORPTION TO SOIL AND SEDIMENT Free cyanide (CN− , HCN) adsorbs weakly on soils and sediment. The cyanide anion can be retained by soils with anion exchange capacity, but in the pH range 4 to 9 of interest for most soils, HCN is the dominant form of cyanide and CN− concentrations are very low. HCN adsorbs weakly or not all to inorganic soil components such as iron oxide [8], aluminum oxide, clay, and sand [9]. However, HCN has been shown to adsorb significantly to soils with appreciable organic carbon content. The magnitude of cyanide adsorption onto soils tested by Chatwin et al. [10] showed excellent correlation with organic carbon content. Higgins and Dzombak [9] further demonstrated the interaction of HCN with organic carbon in experiments with activated carbon and freshwater sediment. They developed an expression relating sorbed HCN concentration, CS , to aqueous phase concentration, Cw , through an organic carbon normalized distribution coefficient Koc (=Kd /foc ). CS = Koc Cw foc = (6.5 L/gs )Cw foc

(5.9)

where CS is in µg/gs , CW is in µg/L, and foc is the fraction of organic carbon in the adsorbent. The experiments upon which this linear relationship is based all involved low concentrations of free cyanide in water (<150 µg/L), which is typical for total cyanide concentrations encountered in environmental contamination scenarios. Adsorption capacities were not determined in the experiments with activated carbon and sediment. Literature data on free cyanide adsorption onto activated carbon have shown an adsorption capacity of about 1 to 2 mg of free cyanide per gram of carbon, while similar tests performed with soil organic carbon have revealed an adsorption capacity of 0.5 mg of free cyanide per gram of carbon [11]. Batch and column tests performed by Alesii and Fuller [12] with various soils yielded significant removal of free cyanide at near-neutral pH values. Soil constituents included kaolin clay, chlorite, gibbsite clay, and iron and aluminum oxides. Based on the laboratory results discussed earlier, it is unlikely that these inorganic constituents would adsorb free cyanide to an appreciable extent. As the soils used in the experiments by Alesii and Fuller were not sterilized and hence biologically active, it is more likely that the free cyanide was removed from the system via biodegradation. Dzombak and Morel [13] estimated equilibrium surface complexation constants for the adsorption of CN− , CNO− , and SCN− on hydrous ferric oxide based on correlations of acidity constants and surface complexation constants fitted to adsorption data for other inorganic ions. The surface complexation reactions and the estimated surface complexation constants for those reactions are given in Table 5.2. Predicted adsorption versus pH curves for 10−4 M solutions of these ions in hydrous ferric oxide suspensions with TOTFe = 10−3 M and ionic strength of 0.01 M are shown in Figure 5.3. While the accuracy of these predictions is uncertain due to the estimated nature of the surface complexation constants, the predictions provide some idea of the expected adsorption behavior based on what has been observed with other inorganic ions. Available data for free cyanide adsorption on mineral surfaces, however, indicates that the free cyanide adsorption in Figure 5.3 is likely to be substantially overpredicted. Free cyanide has been observed to exhibit little to no adsorption on mineral surfaces, including the crystalline iron oxide goethite, across a range of pH [8,9].

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TABLE 5.2 Estimated Surface Complexation Reactions and Constants for Adsorption of CN− , CNO− , and SCN− on Hydrous Ferric Oxide Adsorbing species, A− CN− CNO− SCN−

log K2 (25◦ C, I = 0)a

log K3 (25◦ C, I = 0)b

13.0 8.9 7.0

5.7 1.8 0.1

a SC reaction: ≡FeOH0 + A− + H+ = ≡FeA0 + H O; K . 2 2 b SC reaction: ≡FeOH0 + A− = ≡FeOHA− ; K . 3

Source: Data from Dzombak, D.A. and Morel, F.M.M., Surface Complexation Modeling: Hydrous Ferric Oxide, Wiley-Interscience, New York, NY, 1990 (Table 10.10).

80 70

Percent adsorbed

60 50 40 CN– CNO– SCN–

30 20 10 0 4

5

6

7 pH

8

9

10

FIGURE 5.3 Predicted adsorption of 10−4 M CN− , CNO− , and SCN− on hydrous ferric oxide as a function of pH. Predictions made using surface complexation constants of Dzombak, D.A. and Morel, F.M., Surface Complexation Modeling: Hydrous Ferric Oxide, 1990; see Table 5.2. TOTFe = 0.001 M, I = 0.1 M. Adsorption of CN− is likely overpredicted.

5.1.5 FREE CYANIDE OXIDATION Free cyanide can be oxidized to cyanate, CNO− , or hydrogen cyanate, HCNO, depending on the pH [14]: CN− + H2 O = CNO− + 2H+ + 2e−

(5.10)

HCN + H2 O = HCNO + 2H+ + 2e−

(5.11)

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Cyanate is protonated only at low pH, as the pK a is 3.45 [5]: HOCN = H+ + CNO− ,

pK a = 3.45 at 25◦ C,

I=0

(5.12)

In the oxidation reactions of Equations (5.10) and (5.11), the oxidation state of carbon in CN− is +2, while in CNO− , the oxidation state of carbon is +4. Free cyanide can be oxidized readily by strong oxidants such as chlorine, hypochlorite, ozone, and hydrogen peroxide [15–18]. Under neutral to alkaline conditions, the end product is cyanate (CNO− ), which is relatively nontoxic. The oxidative conversion of CN− to CNO− in alkaline chlorination is often exploited for rapid treatment of free cyanide in water. The general reaction chemistry for alkaline chlorination is as follows [17]: CN− + Cl2 → CNCl + Cl− −

(5.13) +

−

CNCl + 2NaOH → CNO + 2Na + 2Cl + H2 O −

3Cl2 + 2CNO + 6NaOH →

2HCO− 3

−

(5.14) +

+ N2 + 6Cl + 6Na + 2H2 O

(5.15)

As indicated in Equation (5.13), cyanide ion is first converted to cyanogen chloride, CNCl, a highly toxic species. Under alkaline conditions, the CNCl reacts rapidly with OH− to form CNO− , and upon further chlorination the cyanate decomposes to form the completely benign products bicarbonate, HCO− 3 and elemental nitrogen, N2 . In the last step, Equation (5.15), the nitrogen is oxidized, moving from an oxidation state of −3 to zero. Gurol and Bremen [19] studied ozonation of free cyanide. It was found that the ozone molecule, O3 , reacts primarily with the cyanide ion; its reaction with HCN is negligible. Further, it was determined that the presence of free cyanide promotes the formation of free radicals (OH• , HO2 • ), and that free radical reactions as well as direct reaction of the free cyanide with ozone contribute to the overall oxidative destruction of the cyanide. Hence, there are numerous initiators and pathways involved in the oxidation of free cyanide by ozone. Some of the reactions identified by Gurol and Bremen [19] as involved with the ozonation of free cyanide are as follows: O3 + CN− → CNO− + O2

(5.16)

HCN + OH• → HCNO + HO2 •

(5.17)

−

−

CN + OH• → CNO + HO2 •

(5.18)

CN− + OH• → CN• + OH−

(5.19)

CN• + CN• → (CN)2 −

(5.20) −

−

(CN)2 + 2OH → CNO + CN + H2 O

(5.21)

The direct reaction of molecular ozone with the cyanide ion is indicated in Equation (5.16). Other − reactions of ozone with water, specifically OH− , yield the superoxide radical O2 • , which reacts further with ozone to give the hydroxyl radical OH• . The oxidation of cyanide by ozone is rapid and pH dependent [19]. Solutions of several mM of free cyanide were oxidized within 5 to 30 min by ozone, with faster rates at higher pH values where more of the free cyanide was in the form of CN− . The end product of ozonation of free cyanide is cyanate. The cyanate is further oxidized by ozone, but since this is a relatively slow reaction cyanate accumulates in solution until free cyanide is oxidized completely [20]. Free cyanide in the environment is oxidized rapidly by aerobic bacteria, for which it can serve as an energy source, as discussed in Chapters 6 and 23. Thus, abiotic oxidation of free cyanide in the environment is usually of secondary concern.

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Chatwin et al. [10] detected cyanate in effluent from saturated soil columns through which aqueous solutions containing free cyanide were passed. It was hypothesized that the free cyanide was oxidized to cyanate on the surfaces of clay components of the soil, and that the process was enhanced with the addition of copper and nickel to the system. However, since the soils studied were microbiologically active, microbial degradation was likely to have had an equal or possibly greater role in the conversion of the cyanide, a factor not addressed by Chatwin et al. [10]. Based on other work showing very limited to no interaction of free cyanide with mineral surfaces [8], it appears unlikely that abiotic oxidation of free cyanide on mineral surfaces will occur appreciably in natural systems. Free cyanide can also react with and be oxidized by various forms of sulfur, especially poly− sulfides and thiosulfate (S2 O2− 3 ), to form thiocyanate (SCN ). In neutral to alkaline solutions, both polysulfides and thiosulfate are products of oxidation of sulfide. The reactions of polysulfide and thiosulfate with the cyanide ion are as follows [2,21]: Sx S2− + CN− → Sx−1 S2− + SCN−

(5.22)

2− − − S2 O2− 3 + CN → SO3 + SCN

(5.23)

For thiocyanate, the oxidation states of the S, C, and N are −1, +3, and −3, respectively. In the reaction of polysulfide with free cyanide (Equation 5.22), one polysulfide sulfur atom is reduced from its oxidation state 0 to −1, while the cyanide carbon atom is oxidized from +2 to +3 [21]. In the reaction of thiosulfate with free cyanide (Equation 5.23), one thiosulfate sulfur atom changes from oxidation state +2 to +4, while the other thiosulfate sulfur atom is reduced from the +2 to the −1 oxidation state [21]. The rate of thiocyanate formation through reaction of polysulfide and free cyanide is approximately three orders of magnitude greater than through reaction of free cyanide and thiosulfate, depending on pH [21]. Thus, in systems with equal amounts of polysulfide and thiosulfate present, the reaction of free cyanide with polysulfide will be the dominant thiocyanate formation route. The formation of polysulfide through oxidation of sulfide occurs at a slow rate, however, so available polysulfide is often limited [21]. Reaction of thiosulfate with free cyanide thus governs the formation of thiocyanate in many systems.

5.1.6 FREE CYANIDE HYDROLYSIS As discussed in Section 5.1.2, the cyanide ion reacts with water (H+ ) to form HCN, with the protonated species HCN being the dominant form of free cyanide at pH values less than 9.24 at 25◦ C (Figure 5.1). At ambient temperatures, this protonation reaction is the primary reaction of free cyanide with water. Free cyanide can react with molecular water under alkaline conditions and high temperature to yield formate and ammonia: CN− + 2H2 O → HCOO− + NH3

(5.24)

The reaction proceeds at appreciable rates only at high temperatures, and at fast rates at high temperature and pressure, for example, temperatures in the range of 165–180◦ C and pressures of 100–150 psig [22]; see Chapters 20 and 22. Wiegand and Tremelling [23] showed that the rate of free cyanide hydrolysis is very slow at room temperature, increasing about threefold for every 10◦ C rise in temperature. At lower pH values, HCN can also be hydrolyzed, yielding formic acid and ammonia [2]: HCN + 2H2 O → HCOOH + NH3 Under acidic conditions the reaction is also very slow.

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(5.25)

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65

Alkaline hydrolysis has been exploited for treatment of free and complexed cyanide in wastewaters and sludges [22,24]. Alkaline conditions assure that any free cyanide remains dissolved in the form of CN− during the treatment process. Temperatures in the range of 140 to 275◦ C, and pressures up to 900 psig are employed in the alkaline hydrolysis process. This treatment process is discussed in more detail in Chapter 22.

5.2 METAL CYANIDES: AQUEOUS SPECIES Many metals form aqueous complexes with cyanide ion by means of π bonding, which occurs when the participating metal atom donates one or more electrons to CN− , which serves as an electron accepting ligand. These soluble metal–cyanide complexes, represented by a general formula [M(CN)x ]y− , where, M signifies a metal cation, can be classified into weak and strong metal–cyanide complexes, depending upon the strength of the metal–cyanide bonding. Use of vibrational spectroscopy reveals different electronic structures of [M(CN)x ]y− complexes [1]. Depending on the different modes of vibration, a [M(CN)x ]y− species can exist in tetrahedral, square planar, or octahedral forms. These common electronic structures are shown in Figure 5.4.

5.2.1 WEAK METAL–CYANIDE COMPLEXES 5.2.1.1 Formation The cyanide anion can form weak metal–cyanide complexes with many transition metals, the most common among them being cadmium, zinc, silver, copper, nickel, and mercury. Most of these metals fall in Groups IB, IIB, and VIIIB of the periodic table. The metal–cyanide bonds in these complexes are mostly arranged in tetrahedral or square planar forms with relatively weak bonding energy existing between the heavy metal atom and the cyanide ligand as compared to the strong cyanide complexes with iron, cobalt, and platinum. Because weakly-bonded metal–cyanide complexes dissociate under weakly acidic pH conditions (4 < pH < 6), they are commonly termed weak-acid-dissociable (WAD) complexes [15]. Formation data determined by direct thermodynamic measurements are available for complexes of nickel(II), copper(I), silver(I), zinc(II), cadmium(II), and mercury(II) [1,5]. Equilibrium or stability constants have been calculated from standard thermodynamic data for a broad range of metal–cyanide complexes, however [5]. Table 5.3 lists the measured and calculated stability constants compiled by Sehmel [5] for formation of weak metal–cyanide complexes. For comparable reaction stoichiometry, the higher the value of the formation equilibrium constant (K), the greater is the energy of formation and the stability of the metal–cyanide complex.

Fe (III)

N

N N

N C

CC

C

( Fe

Ni2+

N

III)

C

N

C

CC

C N

C

N N

N

[Ni(CN) 4] 2– (tetracyanonickelate)

[Fe(CN) 6] 3– (ferricyanide)

FIGURE 5.4 Common electronic structures of metal–cyanide complexes.

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TABLE 5.3 Equilibrium Constants for Metal–Cyanide Complexes

Formation

of

Reaction Ag+ + CN− + H2 O = AgCN(OH)−

Ag+ + 2CN− = Ag(CN)− 2 Ag+ + 3CN− = Ag(CN)2− 3 Ag+ + 2OCN− = Ag(OCN)− 2 Cd2+ + CN− = CdCN−

Cd2+ + 2CN− = Cd(CN)02 Cd2+ + 3CN− = Cd(CN)− 3 Cd2+ + 4CN− = Cd(CN)2− 4 Cu+ + 2CN− = Cu(CN)− 2 Cu+ + 3CN− = Cu(CN)2− 3 Cu+ + 4CN− = Cu(CN)3− 4 Ni2+ + 2CN− = Ni(CN)02 Ni2+ + 3CN− = Ni(CN)− 3 Ni2+ + 4CN− = Ni(CN)2− 4 Ni2+ + H+ + 4CN− = NiH(CN)− 4 Ni2+ + 2H+ + 4CN− = NiH2 (CN)04 Ni2+ + 3H+ + 4CN− = NiH3 (CN)+ 4 Zn2+ + 2CN− = Zn(CN)02 Zn2+ + 3CN− = Zn(CN)− 3 Zn2+ + 4CN− = Zn(CN)2− 4 Hg(OH)02 + 2H+ + CN− = HgCN+ + 2H2 O Hg(OH)02 + 2H+ + 2CN− = Hg(CN)02 + 2H2 O Hg(OH)02 + 2H+ + 3CN− = Hg(CN)− 3 + 2H2 O 0 + − Hg(OH)2 + 2H + 4CN = Hg(CN)2− 4 + 2H2 O Hg(OH)02 + 2H+ + 2CN− + Cl− = Hg(CN)2 Cl− + 2H2 O Hg(OH)02 + 2H+ + 3CN− + Cl− = Hg(CN)3 Cl2− + 2H2 O Hg(OH)02 + 2H+ + 3CN− + Br − = Hg(CN)3 Br 2− + 2H2 O

Selected

Weak

log K (at 25◦ C, I = 0) −0.56 20.38 21.40 5.00 5.32 10.37 14.83 18.29 24.03 28.65 30.35 14.59 22.63 30.13 36.75 41.46 43.95 11.07 16.05 16.72 24.17 40.65 44.40 47.41 40.37 43.83 44.94

Source: Data from Sehmel, G.A., Cyanide and antimony thermodynamic database for the aqueous species and solids for the EPA-MINTEQ geochemical code, PNL-6835, Pacific Northwest Laboratory, Richland, WA, 1989, (Table 5).

The equilibrium constants compiled by Sehmel [5] were selected and included in Table 5.3 rather than those reported in some other compilations, for example, the work of Beck [25] and Martell et al. (1993), because the constants reported by Sehmel were calculated in a consistent manner using the most current thermodynamic data from the U.S. National Bureau of Standards [26,27]. The constants reported in Beck [25] were calculated using older (1952) NBS thermodynamic data [28]. As shown by Gilgore-Schnorr and Dzombak [29], the key difference is in the value used for the partial molar entropy, So , of the cyanide ion CN− , for which the 1952 [28] value of 28.2 cal K −1 mol−1 was revised in 1965 [27] to 22.5 cal K−1 mol−1 , a value retained in the 1982 (and most current) thermodynamic data compilation [26]. The work of Sehmel [5] was performed for the USEPA, which incorporated the metal–cyanide complexation constants in the thermodynamic database of the general chemical

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100 Zn+2 Zn(CN)4+2

ZnT = 2.00
10–5 M

80

Percent Zn species

CNT = 2.00
10–3 M

Zn(CN)02

I = 0.1 M

60

PCO2 = 10–3.5 atm 40

Zn(CN)3–

20

Zn(CN)1+

0 3

4

5

6 pH

7

8

9

FIGURE 5.5 Calculated aqueous speciation of zinc(II) in the presence of excess cyanide. ZnT = 10−4.70 M, CNT = 10−2.70 M, PCO2 = 10−3.5 atm, I = 0.1 M. (Source: Theis, T.L. and West, M.L., Environ. Technol. Lett., 7, 309, 1986.)

equilibrium program MINTEQ [30,31]. Several other general chemical equilibrium programs employ the MINTEQ thermodynamic database, such as MINEQL+ [32]. An example of zinc speciation in aqueous solution calculated by Theis and West [8] is presented in Figure 5.5. (Note: Equilibrium constants from Sehmel [5] were not used in calculating the speciation plots of Figure 5.5; the plots are for illustrative purposes.) In the system shown, the total amount of free cyanide added to the system (2 × 10−3 M) is in excess of the amount of zinc present (2 × 10−5 M). Thus, cyanide species dominate the speciation of zinc above pH 6.2. At lower pH values, H+ outcompetes Zn2+ for complexation with CN− as H+ becomes more abundant and HCN forms. Soluble zinc hydroxide complexes also form, at higher pH, but in this system CN− outcompetes OH− in the pH range shown. Concentrations of dissolved zinc hydroxide species are very small and their influence on zinc speciation is not significant. As is evident in Figure 5.5, different zinc cyanide species predominate in different pH regions, with Zn(CN)02 the dominant form from pH 6.2 to 6.7, and Zn(CN)2− 4 being most abundant above pH 6.7. The kinetics of metal–cyanide complex formation can be slow [1,33]. For example, Broderius [33] showed that Ni(CN)2− 4 formation took three days to reach equilibrium (0.5 to 6.5 ppm CNT ), while formation of copper cyanide complexes took over 100 days to reach equilibrium. The slow rate of formation of metal–cyanide complexes, and the potential for oxidation of many weak metal–cyanide complexes by atmospheric oxygen, makes it difficult to measure the equilibrium constants for these complexes [1]. 5.2.1.2 Dissociation The dissociation properties of weak metal–cyanide complexes in aqueous solutions depend on their stability constants, pH, temperature and the redox potential of the solution. In general, metal–cyanide complex dissociation may be described by + − M(CN)n− x = M + xCN

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(5.26)

Cyanide in Water and Soil

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where M+ is a monovalent metal cation, x is the number of cyanide groups, and n is the ionic charge of the metal–cyanide complex. The stability constants given in Table 5.3 provide an indication of the propensity of weak acid metal–cyanide complexes in solution to dissociate. The higher the value for K for reactions with the same stoichiometry, the more limited the dissociation into free cyanide (CN− ). Because of the labile nature of weak metal–cyanide complexes, mildly acidic conditions (pH ≈ 4 to 6) can result in the dissociation of many of these complexes, especially nickel and zinc. For this reason, the weak-acid-dissociable cyanide analytical method [15] employs a pH 4.5 buffer while distilling the aqueous sample at 125◦ C for 1.5 h to break down these complexes completely and quantify the amount of cyanide associated with them (see Chapter 7).

5.2.1.3 Adsorption on Soil and Sediment Weak metal–cyanide complexes can adsorb on common soil and sediment components such as iron, aluminum, silicon, and manganese oxides, and clays, which in most systems will inhibit their aqueous transport [8,12,34]. However, complexation of metals by cyanide can also serve to hold them in solution, inhibiting their adsorption and retention. The enhancement or inhibition of adsorption depends on the metal–cyanide species, the adsorbent, and the solution conditions. Theis and Richter [34] studied the adsorption of the predominant nickel–cyanide anion Ni(CN)2− 4 on silicon dioxide, SiO2 (s) and the iron oxide goethite, FeOOH(s). Batch adsorption experiments were conducted in 0.01 M NaClO4 aqueous solutions containing 10−4.77 M total nickel (NiT ) and amounts of total free cyanide (CNT ) of 10−5 , 10−4 , and 10−3 M. Calculated plots of the equilibrium distribution of nickel species as a function of pH in aqueous solution with no solids present are given in Figure 5.6. (Note: Equilibrium constants from Sehmel [5] were not used in calculating the speciation plots of Figure 5.6; the plots are for illustrative purposes.) As may be seen there, Ni(CN)2− 4 is predicted to dominate nickel speciation at pH < 5.5 in the system with CNT = −4 10 M, and at pH < 4.5 in the system with CNT = 10−3 M. A speciation diagram for the system with CNT = 10−5 M is not shown, but with NiT in excess of CNT in this system Ni2+ and NiOH+ are the dominant forms of nickel and a maximum of about 10% of the nickel becomes bound to cyanide. Results of the batch adsorption experiments conducted with SiO2 (s) are shown in Figure 5.7. The adsorption of nickel observed in the system with CNT = 10−5 M is very similar to the adsorption of nickel on SiO2 (s) in the absence of cyanide. For the series of batch experiments with CNT = 10−4 and 10−3 M, however, adsorption of nickel is inhibited. These data indicate that the nickel–cyanide species Ni(CN)2− 4 , which dominates nickel speciation in both systems, has no affinity for the SiO2 (s) surface. The SiO2 (s) surface is negatively charged for pH > 2, so electrostatic repulsion of the negative Ni(CN)2− 4 species is in part responsible for the absence of surface binding. For the most part, however, it is the presence of cyanide that inhibits adsorption of nickel. Figure 5.8 shows data for similar sets of batch adsorption experiments performed with FeOOH(s). Once again, the adsorption of nickel observed in the system with CNT = 10−5 M is very similar to the adsorption of nickel on FeOOH(s) in the absence of cyanide. In the systems with CNT = 10−4 and 10−3 M, adsorption of nickel is enhanced at lower pH values, and inhibited at higher pH values. The FeOOH(s) surface is positively charged up to about pH 6, or even higher, so electrostatic attraction of Ni(CN)2− 4 explains in part its adsorption at lower pH values. Electrostatic attraction alone is not sufficient to explain the extent of removal observed, however. Through surface interaction modeling, Theis and Richter [34] demonstrated that Ni(CN)2− 4 must bond at specific surface sites on goethite, in surface complexation reactions that involve high free energies of interaction. They proposed the formation of a goethite–cyanide–metal surface complex via a surface complexation reaction: 2− + ≡Fe2 (OH)2+ 2 + Ni(CN)4 + 2H = ≡Fe2 –(CN)2 –Ni–(CN)2 + 2H2 O

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(5.27)

Physical–Chemical Properties and Reactivity

69

(a) 100 Ni2+

Ni(CN)42–

Percent total Ni

80

60

CNT = 10–4 M

40

20

0 3

4

5

6

7

8

9

10

pH (b) 100 Ni(CN)42–

Ni2+

CNT = 10–3 M

Percent total Ni

80

60

40

20

0 3

4

5

6

7

8

9

10

pH

FIGURE 5.6 Theoretical distribution of nickel in the presence of (a) 10−4 M cyanide (NiT = 10−4.77 M, I = 0.01 M), and (b) 10−3 M cyanide (NiT = 10−4.77 M, I = 0.01 M). (Source: Reprinted with permission from Theis, T.L. and Richer, R.O., Particulates in water, 189, 73, 1980. Copyright 1980 American Chemical Society.)

where ≡Fe2 (OH)2+ 2 is a surface hydroxyl site on the surface of goethite in aqueous suspension, and ≡Fe2 –(CN)2 –Ni–(CN)2 is the surface species formed by adsorption of Ni(CN)2− 4 on the goethite. The uptake of H+ shown in the reaction occurs commonly in adsorption of inorganic anions on oxides, and is related to the commonly observed pH dependence for anion adsorption: maximum adsorption at lower pH and decrease in adsorption with increasing pH [13]. Formation of a metal– ligand–metal ternary surface complex as shown in the reaction of Equation (5.27) has been proposed for other metal–ligand systems [35–38]. Theis and West [8] studied the adsorption of cadmium, copper, and zinc divalent cations and their metal–cyanide complexes on goethite in aqueous suspensions. Some typical results for adsorption of Cd2+ , Cu2+ , and Zn2+ in the absence of cyanide are presented in Figure 5.9. All of the experiments were conducted with total metal concentration of approximately 2 × 10−5 M and with 0.6 g/L

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70

100

90

80

Percent nickel removed

70

60

50

40

30

20

10

0 3

4

5

6

7 pH

8

9

10

11

FIGURE 5.7 Nickel adsorption as a function of pH in the presence of silicon dioxide and Cyanide. NiT = 10−4.77 M, I = 0.01 M, SiO2 = 29.41 g/L. (
) CNT = 10−5 M, () CNT = 10−4 M, () CNT = 10−3 M, ( ) CNT = 0. (Source: Reprinted with permission from Theis, T.L. and Richer, R.O., Particulates in water, 189, 73, 1980. Copyright 1980 American Chemical Society.)

•

FeOOH(s). The pH adsorption edge plots shown in Figure 5.9 exhibit the typical characteristics for cation adsorption on metal oxides: an increase from 0 to 100% adsorbed with increasing pH. Batch adsorption experiments conducted with free cyanide showed no adsorption of the free cyanide on goethite for any pH from 3 to 11 (data not shown). Experiments with free cyanide added in excess of the metal concentrations were also performed. Results for the three metals are given in Figure 5.10. At lower pH values, adsorption of the cadmium, copper, and zinc was unaffected by the free cyanide as may be seen by comparison with Figure 5.9. Above pH 6.5 to 7.0, however, adsorption of the metals was inhibited by the presence of the cyanide. At the higher pH values, metal–cyanide complexes dominate the speciation of the metals (e.g., see the aqueous phase speciation diagram for zinc in Figure 5.5). The data in Figure 5.10 indicate that the cadmium–, copper–, and zinc–cyanide species have no affinity for the goethite surface at neutral to alkaline pH values. The examples presented in this section demonstrate that some weak metal–cyanide complexes can adsorb on soils and soil components under some conditions, but the extent of adsorption depends strongly on the particular metal–cyanide species, mineral adsorbent, and solution conditions. Solution pH is an especially important governing parameter, as is the case for adsorption of all ions on oxidic minerals [13]. The data presented also demonstrate that the presence of free cyanide in

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100

90

80

Percent nickel removed

70

60

50

40

30

20

10

0 3

4

5

6

7

8

9

10

11

pH

FIGURE 5.8 Nickel adsorption as a function of pH in the presence of iron oxide (goethite) and cyanide. NiT = 10−4.77 M, I = 0.01 M, α-FeOOH = 0.59 g/L. (
) CNT = 10−5 M, () CNT = 10−4 M, () CNT = 10−3 M, ( ) CNT = 0. (Source: Reprinted with permission from Theis, T.L. and Richer, R.O., Particulates in water, 189, 73, 1980. Copyright 1980 American Chemical Society.)

•

a systems with metals, leading to the formation of metal–cyanide complexes, can result in enhanced or reduced adsorption of the metals. The metal–cyanide complexes may interact with the surface to a greater or lesser extent than the metals alone. An interrelated, complex group of factors governs metal–cyanide species adsorption, and it is diffcult to form generalizations. 5.2.1.4 Oxidation Weak metal–cyanide complexes generally are readily oxidized by oxidizing agents such as chlorine or ozone. The more strongly bonded complexes in the WAD category, such as nickel, silver, and mercury cyanide complexes, oxidize more slowly [15]. The more weakly-bonded complexes, including those of cadmium, copper, and zinc, decompose rapidly in the presence of oxidizing agents. As discussed in Section 5.1.5, and in more detail in Chapter 20, alkaline chlorination is the most common approach used to treat waters bearing free cyanide. A number of weak metal–cyanide complexes are also readily oxidized in this process. In order to identify the fraction of measured total cyanide, which includes metal–cyanide complexes and free cyanide, that is treatable by alkaline

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100 90 80 Percent removed

70 60 50

Cd+2 = 2.08
10–5 M Cu+2 = 1.92
10–5 M

40

Zn+2 = 2.41
10–5 M

30

α −FeOOH = 0.6 g/L I = 0.1 M

20 10 0 3

4

5

6

7

8

9

10

11

pH

FIGURE 5.9 Adsorption of Cd, Cu, and Zn on goethite. I = 0.01 M, α-FeOOH = 0.6 g/L. () CuT = 10−4.72 M, () ZnT = 10−4.62 M, (
) CdT = 10−4.68 M. (Source: Theis, T.L. and West, M.L., Environ. Technol. Lett., 7, 309, 1986.) 100 ZnT = 1.73
10–5 M/ CNT = 2.00
10–3 M CuT = 1.95
10–5 M/ CNT = 1.00
10–4 M CdT = 1.88
10–5 M/ CNT = 2.00
10–3 M

90 80 Percent removed

70 60 a−FeOOH = 0.6 g/L I = 0.1M

50 40 30 20 10 0 3

4

5

6

7

8

9

10

11

pH

FIGURE 5.10 Effect of cyanide on adsorption of Cd, Cu, and Zn on goethite. I = 0.01 M, α-FeOOH = 0.6 g/L. (
) CdT = 10−4.73 M, CNT = 10−2.70 M; () ZnT = 10−4.76 M, CNT = 10−2.70 M; () CuT = 10−4.71 M, CNT = 10−4.00 M. (Source: Theis, T.L. and West, M.L., Environ. Technol. Lett., 7, 309, 1986.)

chlorination, an analytical measurement known as “cyanide amenable to chlorination” [15] has long been employed. The method involves measurement of total cyanide on samples with and without treatment by chlorination, with the difference giving the amount of cyanide in the sample amenable to chlorination (Chapter 7). While the CATC method has limitations, as discussed in Chapter 7, the existence of the method speaks to the facile oxidation of a number of weak metal–cyanide complexes. In some cases, the presence of weak metal–cyanide complexes can enhance the rate of free cyanide decomposition through catalysis by the metal. This has been demonstrated for copper cyanide complexes [20]. Gurol and Holden [20] studied the effect of copper(I) on the removal of free cyanide by

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ozone in alkaline solution. They performed experiments in solutions at pH 11.5 in systems with an excess of free cyanide over copper(I), giving Cu(CN)3− 4 as the dominant copper species. They found that the presence of copper increased the rate of free cyanide oxidation significantly. Comparison of initial rates of cyanide disappearance for systems with and without copper indicated a fivefold higher rate for the system with copper cyanide species. Further investigations revealed that the observed enhancement was likely due a very fast, independent oxidation–reduction reaction between Cu(I) and free cyanide. The following reaction sequence was proposed: 2− − 2Cu(CN)3− 4 + 2O3 → 2Cu(CN)3 + 2CNO + 2O2

(5.28)

− + 2Cu(CN)2− 3 + O3 + 2H → 2Cu(CN)3 + O2 + H2 O

(5.29)

− 2Cu(CN)− 3 → 2Cu(CN)2 + (CN)2

(5.30)

2Cu(CN)− 2

(5.31)

+ 4CN

−

= 2Cu(CN)3− 4 − −

(CN)2 + 2OH− → CN + CNO + H2 O

(5.32)

The reaction in Equation (5.28) represents the direct oxidation of cyanide to cyanate by ozone. In Equation (5.29), Cu(I) is oxidized to Cu(II) in a fast reaction. The Cu(II) species subsequently oxidizes cyanide to cyanogen (C2 N2 ), being reduced back to Cu(I) in the process (Equation [5.30]). An equilibrium between the copper(I) cyanogen species and Cu(CN)3− 4 is rapidly established (Equation [5.31]). In the last step, Equation (5.32), cyanogen goes through a disproportionation reaction to yield free cyanide and cyanate. The net reaction from the above sequence is thus as follows: 3− − − 2Cu(CN)3− 4 + 3CN + 3O3 → 2Cu(CN)4 + 3CNO + 3O2

(5.33)

Thus, the oxidation of 3 mol free cyanide requires 2 mol of ozone and produces 3 mol of cyanate, as would be expected, but the rate of the reaction is much accelerated due to the presence of the Cu(I).

5.2.2 STRONG METAL–CYANIDE COMPLEXES 5.2.2.1 Formation The cyanide anion can form strong complexes with a number of transition heavy metals, the most notable among them are cobalt, platinum, gold, palladium, and iron. Most of these metals fall in Groups IB, IIB, and VIII of the periodic table. As iron is by far the most abundant of these elements in the environment and in process waters, iron–cyanide complexes are the strong metal–cyanide complexes of greatest interest. Gold–cyanide complexes are of great commercial interest, as the strength of the gold–cyanide bond is exploited in hydrometallurgical gold mining for aqueous extraction of gold from ores (see Chapters 4 and 26). The metal–cyanide bonds in these complexes are mostly arranged in tetrahedral or octahedral forms with strong bonding energy existing between the heavy metal atom and the cyanide ligand [1]. Because they can only dissociate under strongly acidic pH conditions (pH < 2), they are referred to as strong acid dissociable complexes, or simply as strongly-complexed cyanide [15]. As some of these species are formed very slowly [1,33], it is difficult to determine the equilibrium formation constants. Formation data determined by direct thermodynamic methods are available only for complexes of gold(I), and palladium(II) [1]. For other metals, like iron, electron transfer between complex ions of the element in the +2 and +3 oxidation states is rapid enough such that the ratio of the formation constants can be determined from measurement of redox potentials. This ratio can then be combined with standard enthalpy and entropy change measurements for the formation reaction of interest.

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TABLE 5.4 Equilibrium Constants for Formation of Selected Strong Metal–Cyanide Complexes Reaction Ba2+ + Fe2+ + 6CN− = BaFe(CN)2− 6

Ba2+ + Fe3+ + 6CN− = BaFe(CN)− 6 Ca2+ + Fe2+ + 6CN− = CaFe(CN)2− 6 Ca2+ + Fe3+ + 6CN− = CaFe(CN)− 6 2Ca2+ + Fe2+ + 6CN− = Ca2 Fe(CN)06 Ca2+ + H+ + Fe2+ + 6CN− + e− = CaHFe(CN)2− 6 Fe2+ + 6CN− = Fe(CN)4− 6 Fe2+ + H+ + 6CN− = HFe(CN)3− 6 Fe2+ + 2H+ + 6CN− = H2 Fe(CN)2− 6 Fe3+ + 6CN− = Fe(CN)3− 6 2Fe2+ + 6CN− = Fe2 (CN)06 2K+ + Fe2+ + 2H+ + 6CN− = K2 H2 Fe(CN)06 3K+ + Fe2+ + H+ + 6CN− = K3 HFe(CN)06 K+ + Fe2+ + 6CN− = KFe(CN)3− 6 2K+ + Fe2+ + 6CN− = K2 Fe(CN)2− 6 K+ + Fe2+ + H+ + 6CN− = KHFe(CN)2− 6 Li+ + Fe2+ + 6CN− = LiFe(CN)3− 6 2Li+ + Fe2+ + 6CN− = Li2 Fe(CN)2− 6 Li+ + Fe2+ + H+ + 6CN− = LiHFe(CN)2− 6 Mg2+ + Fe3+ + 6CN− = MgFe(CN)− 6 Mg2+ + Fe2+ + 6CN− = MgFe(CN)2− 6 2+ + 6CN− = NH Fe(CN)3− NH+ 4 4 + Fe 6 2+ + 6CN− = (NH ) Fe(CN)2− + Fe 2NH+ 4 2 4 6 + 2+ + 6CN− = NH Fe(CN)2− NH+ 5 4 + H + Fe 6 Na+ + Fe2+ + 6CN− = NaFe(CN)3− 6 2Na+ + Fe2+ + 6CN− = Na2 Fe(CN)2− 6 Na+ + Fe2+ + H+ + 6CN− = NaHFe(CN)2− 6 Sr 2+ + Fe3+ + 6CN− = SrFe(CN)− 6 Tl+ + Fe2+ + 6CN− = TlFe(CN)3− 6

log K (at 25◦ C, I = 0) 49.40 55.44 49.69 55.47 51.00 52.71 45.61 50.00 52.45 52.63 56.98 52.31 50.22 48.12 48.98 51.47 47.69 48.53 51.22 55.39 49.43 48.07 48.87 51.40 47.99 48.74 51.43 55.62 48.75

Source: Data from Sehmel, G.A., Cyanide and antimony thermodynamic database for the aqueous species and solids for the EPA-MINTEQ geochemical code, PNL-6835, Pacific Northwest Laboratory, Richland, WA, 1989, (Table 5).

Table 5.4 lists the equilibrium constants for the reversible formation of iron–cyanide complexes, which are of primary interest with respect to cyanide in the environment. The constants reported are from the compilation by Sehmel [5], which was selected for the reasons discussed in Section 5.2.1.1. Among all the iron–cyanide complexes, the most commonly occurring are ferrocyanide, Fe(CN)4− 6 , 3− where iron is the +2 oxidation state, and ferricyanide, Fe(CN)6 , where iron is in the +3 oxidation state. Another iron–cyanide complex only recently identified and not presented in Table 5.4, with a chemical formula, Fe(CN)5 NHCH4− 3 , has been found to dominate groundwater at many former manufactured gas plant sites [39].

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25 20

Water oxidized

15 10 pE

HCN

Fe(CN)6 (tot)

5 0 –5

CN–

–10

Water reduced

–15 0

2

4

6

8

10

pH

FIGURE 5.11 Predominance diagram for dissolved cyanide species in equilibrium with hydrous ferric oxide at T = 25◦ C, as calculated with MINEQL+ (Schecher et al., 1998) using the reactions in Equation (5.1) and Table 5.4. TOTCN = 0.6 mM, TOTFe = 0.5 mM, TOTK = 0.4 mM, TOTNa = 0.06 mM, TOTCl = 0.06 mM, and I = 0.06 M NaCl. (Source: Ghosh, R.S. et al., Water Environ. Res., 71, 1205, 1999.)

In soils and aquifer systems where iron is ubiquitous, the aqueous speciation of cyanide is influenced significantly by reactions with iron dissolved from iron oxides [40,41]. Equilibrium with hydrous ferric oxide, the common amorphous iron oxide, typically is important because Fe(OH)3 (s) serves as the source of iron that becomes dissolved, which in turn regulates the cyanide speciation. Figure 5.11 presents a species predominance diagram for dissolved cyanide species in a system in equilibrium with hydrous ferric oxide. The diagram was calculated with MINEQL+ [32] using the reactions and equilibrium constants in Equation (5.1) and Table 5.4, and in the MINEQL+ thermodynamic database for the iron dissolution, hydrolysis, and redox reactions. In the area denoted “Fe(CN)6 (tot),” cyanide is predicted to exist at equilibrium primarily as the iron cyanide species 4− Fe(CN)3− 6 (oxic conditions) or Fe(CN)6 (anoxic conditions). In the remaining area HCN is the predominant form of dissolved cyanide, except for a small region at pH > 9.2, the pK a for HCN, above which CN− dominates free cyanide speciation. 5.2.2.2 Dissociation As indicated in the equilibrium species predominance diagram of Figure 5.11, iron–cyanide complexes require acidic conditions to dissociate and form free cyanide. It is important to remember, however, that the species distribution shown in Figure 5.11 reflects equilibrium conditions. Actual species distributions for systems with iron present are strongly governed by kinetics. Dissociation of iron–cyanide complexes in the dark is very slow [42]. Like weak metal–cyanide complexes, the dissociation properties of iron–cyanide complexes in aqueous solutions are functions of their stability constants, pH, temperature, and redox potential of the solution. Meeussen et al. [42] ◦ studied the dissociation of ferrocyanide, Fe(CN)4− 6 , in 1 mM solutions in the dark at 15 C. Based on the results, Meeussen et al. [42] projected half-lives ranging from 1 year under reducing conditions (pE ≈ 5) at pH 4 to 1000 years at the same pH under oxidizing conditions (pE ≈ 10). In these laboratory experiments, maximum decompostion rates were observed at low pH and pE. Actual decomposition rates in the environment could be quite different, for example, through enhancement by catalysts. Nevertheless, the results of Meeussen et al. demonstrate the high degree of stability of

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(a)

(b)

k 6

[Fe(CN)6]4–

Ic

k -6

k r

[Fe(CN)5H2O]3– + CN–

kf

[Fe(CN)6]4*

Kd HCN + OH–

approx. 0.02 [Fe(CN)5H2O]2–+ CN–

h [Fe(CN)6]3– hn [Fe(CN)6]3–* approx. 0.98 [Fe2(CN)10]5– ? [Fe2(CN)10]5–

OH– [Fe(CN)5(OH)]3– h hn … [Fe2(CN)10]6–

Fe(OH)3

FIGURE 5.12 Ferro- (a) and ferricyanide (b) photodissociation reaction pathways. (Sources: Information from: Gaspar, V. and Beck, M.T., Polyhedron, 2, 387, 1983; Fuller, M.W., Aust. J. Chem., 39, 1411, 1986.)

iron–cyanide complexes and the importance of considering kinetics when evaluating dissociation of these complexes. Since strong acid conditions are required to dissociate these complexes, the total cyanide analytical method, which is designed to recover strong metal–cyanide complexes in addition to free cyanide and weak metal–cyanide complexes, employs strong acid pH conditions (pH = 1.5) and heat (125◦ C, for 2 h) to achieve dissociation of all metal–cyanide complexes present ([15,43]; Chapter 7). While ferro- and ferricyanide complexes are quite stable in the dark, they can dissociate rapidly when exposed to light [42,44–46]. For example, in experiments with 1.28 µM ferrocyanide solutions at pH 12 exposed to diffuse daylight, Meeussen et al. [42] observed an initial decomposition rate of approximately 8% per hour. Light in the ultraviolet (UV) range (wavelengths less than 420 nm) is responsible for the photolysis of ferro- and ferricyanide [45]. Some proposed photodissociation pathways for ferro- and ferricyanide are shown in Figure 5.12. The photoactivated dissociation rate depends on light intensity, light wavelength, temperature, presence of catalysts, and other parameters [45,47–49], but the results of Meeussen et al. [42] demonstrate clearly that photodissociation of iron–cyanide complexes upon exposure to natural light can be very rapid. Based on photolysis experiments with ferro- and ferricyanide, Broderius and Smith [45] estimated mid-day half-lives (for mid-summer at the latitude and climatic conditions of St. Paul, MN) for 25 to 100 µg/L concentrations of these species to be 18 and 64 min, respectively. Photolytic degradation of ferro- and ferricyanide follows approximately first-order kinetics [45,49], at least initially, but the rate slows as free cyanide accumulates in solution [49]. While some differences in the rates of photolysis of hexacyanoferrate have been observed at different pH values [45,49], Kuhn and Young [49] found no consistent pattern of initial rate coefficient dependence on pH in studies on solutions at pH 4 to 12. The presence of natural organic matter, or other photoreactive substances in water, can significantly decrease the rate of hexacyanoferrate photolysis [49]. The rate of photochemical dissociation in natural waters is dependent on various environmental factors, including free cyanide content of the solution, sunlight intensity, temperature, turbidity, and depth of the water column [45,50]. In many surface waters, significant photolysis will occur only in the top 50 to 100 cm of the water column where sunlight intensity is sufficient, providing opportunity for dilution of any free cyanide produced [45]. Free cyanide could possibly be undetectable or shortlived [51]. There are scenarios, however, where sunlight intensity may be uniform across the entire water column, such as in shallow ponds and in surface runoff. Hexacyanoferrate contamination of the latter can occur, for example, through the spreading of road salt containing iron cyanide as an anticaking agent [52], that is, the commonly used “blue salt.” More details on the photolysis of iron cyanide species and the role of photolysis in fate and transport of hexacyanoferrate in surface waters are provided in Chapter 9. 5.2.2.3 Adsorption on Soil and Sediment Like the weak metal–cyanide complexes (Section 5.2.1.3), strong complexes such as ferrocyanide, 3− Fe(CN)4− 6 , and ferricyanide, Fe(CN)6 , can adsorb on common soil and sediment components

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Percent Fe(CN)–4 6 adsorbed

100

75

50

3.3 mg g–1 1.7 mg g–1

25

0.83 mg g–1 0.50 mg g–1 0 5

6

7

8

9

10

pH

FIGURE 5.13

Ferrocyanide adsorption on alumina at different adsorbate/adsorbent ratios (mg Fe(CN)4− 6 as

CN per g γ-Al2 O3 (s)). 10−5.19 M Fe(CN)4− 6 and various γ-Al2 O3(s) solid doses in 100 mL 0.01 M NaCl solution. (Source: Bushey, J.T. and Dzombak, D.A., J. Coll. Int. Sci., 272, 46, 2004, John Wiley & Sons. Reproduced with permission.)

such as iron, aluminum, and manganese oxides, and clays [8,12,53]. Adsorption of metal–cyanide complexes occurs through a combination of electrostatic attraction and surface complexation [8]. As most strong metal–cyanide complexes are anionic, they can be substantially adsorbed onto soils with high anion exchange capacity. Solution conditions, especially pH, also affect the extent of adsorption of metal–cyanide complexes in aqueous systems. Alessi and Fuller [12] conducted laboratory column mobility tests in which ferricyanide solution was passed through five different soils of varying physical and chemical characteristics. Based on these tests, it was concluded that soil properties, such as low pH (pH < 5), free iron oxide content, and kaolin, chlorite, and gibbsite type clay (high anion exchange capacity) material increased adsorption of iron–cyanide complexes to soil material. Conversely, soils, sediments, and aquifer materials dominated by sand or other components with high cation exchange capacities tend to be weaker adsorbents for iron–cyanide complexes [40]. Mobility tests performed in fixed-bed columns packed with sand-dominated aquifer material and ferrocyanide-contaminated site groundwater as the mobile phase revealed minimal interaction between the dissolved ferrocyanide complexes and site sand [40]. In a mobility test performed by Ghosh et al. [40], ferrocyanide was observed to break through the column in one pore volume, similar to transport of a conservative tracer. It has been demonstrated in a number of studies [8,53,54] that aluminum and iron oxides, two very common and surface-reactive components of soils and sediments [6,55], can adsorb iron–cyanide species significantly, especially at lower pH values (<7). Bushey and Dzombak [53] studied the equilibrium adsorption of ferrocyanide on the aluminum oxide, γ-Al2 O3 , and found that adsorption increased as pH decreased (Figure 5.13). At pH < 7 and for all lower adsorbate/adsorbent ratios (given as mg Fe(CN)4− 6 as CN per g γ-Al2 O3 ) examined, removal of ferrocyanide from solution was essentially complete. As adsorbate/adsorbent ratio was increased, reflecting a greater loading of Fe(CN)4− 6 relative to available solid mass and surface area in the system, the percent 4− of the total Fe(CN)4− 6 adsorbed decreased. The adsorptive characteristics of Fe(CN)6 shown

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100 90 a–FeOOH = 0.6 g/L

Percent Fe(CN)63– adsorbed

80

I = 0.1 M KNO3

70

2 × 10–5M Fe(CN)6–3

60 50 40 30 20 10 0 3

4

5

6

7 pH

8

9

10

11

FIGURE 5.14 Ferricyanide adsorption on goethite. I = 0.01 M, α-FeOOH = 0.6 g/L, 10−4.70 M Fe(CN)3− 6 . (Source: Theis, T.L. and West, M.L., Environ. Technol. Lett., 7, 309, 1986.)

in Figure 5.13 are similar to those observed generally for adsorption of anionic species on metal oxides [13]. Theis and West [8] studied ferricyanide, Fe(CN)3− 6 , adsorption on goethite (α-FeOOH). Based on their work, it is evident that ferricyanide can adsorb substantially onto the goethite surface under acidic conditions (see Figure 5.14). Rennert and Mansfeldt [54] also studied the adsorption of ferricyanide, as well as ferrocyanide, on goethite. Ferrocyanide adsorbed to a somewhat greater extent than ferricyanide, suggesting that ferrocyanide could be less mobile than ferricyanide in a soil system. Further, Rennert and Mansfeldt [56] investigated the effect of sulfate, an adsorbing anion that competes for adsorption sites on oxides [13], on the adsorption of the iron–cyanide complexes. Over the studied range of pH 3.5 to 8, iron–cyanide complex adsorption was strongly dependent on sulfate concentration and vice versa. iron–cyanide complex adsorption was decreased by the presence of the sulfate, especially at lower pH values. Another study by Rennert and Mansfeldt [57] evaluated the influence of different soil properties on iron–cyanide complex adsorption in soils. This study concluded that adsorption of both iron– cyanide complexes is dependent upon soil organic carbon content, clay, oxalate-extractable Fe (free iron) and oxalate-extractable Al. The latter two parameters are indirect measures of the iron and aluminum oxide content of the soil. For soils with low organic carbon content (<10 g/kg), the adsorption of ferro- and ferricyanide was found to be governed by pH and clay content. For soils with high organic carbon content, clay and oxalate-extractable Al, and pH, were found to regulate adsorption behavior for ferricyanide, whereas ferrocyanide adsorption was regulated only by oxalateextractable Al and pH under the same soil conditions. The findings demonstrated that soil organic matter can have an important role in enhancing the adsorption of both iron–cyanide complexes, possibly by reaction between the iron–cyanide nitrogen and reactive functional groups of surface organic matter (e.g., quinone). 5.2.2.4 Oxidation–Reduction Ferricyanide can be readily reduced to ferrocyanide by a variety of reducing agents. The 4− Fe(CN)3− 6 /Fe(CN)6 redox couple has been used to oxidize phenolic compounds, such as resorcinol,

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for example [58]: 4− 2C6 H6 O2 + 2OH− + 2Fe(CN)3− 6 → C12 H10 O4 + 2Fe(CN)6 + 2H2 O

(5.34)

Ferricyanide has also been used as a redox titrant to investigate oxidation–reduction properties of organic compounds and mixtures of organic compounds, including humic acids [59]. For example, ferricyanide has been used to study the one-electron transfer of hydroquinone at low pH: 4− + QH2 + 2Fe(CN)3− 6 → Q + 2H + 2Fe(CN)6

(5.35)

where QH2 and Q represent hydroquinone and quinone species, respectively [59]. 3− Ferrocyanide, Fe(CN)4− 6 , can be oxidized to ferricyanide, Fe(CN)6 , by molecular oxygen in the dark, but the kinetics of this reaction are slow. Asperger et al. [60] determined the first order rate constant for this reaction to be 10−4 sec−1 in the dark at 40◦ C and pH 4.5. In the absence of UV light, only very strong oxidants like ozone, persulfate, and permanganate can oxidize ferrocyanide ion, in acidic solutions, to ferricyanide [61]. It is difficult to oxidize ferrocyanide under neutral to alkaline pH conditions without UV light. Under oxidant concentrations in excess of stoichiometric requirements, however, ferrocyanide would decompose to CO2 and other benign end products at a very slow rate. Gurol and Holden [20] showed, for example, that oxidation of one mole of iron-complexed cyanide to CO2 required in excess of 30 mol of ozone. Similarly, the presence of excess permanganate can decompose any ferricyanide ion formed from ferrocyanide to CO2 , ferric oxide, and nitrogen oxides [61]. Further discussion of ambient temperature oxidation of cyanide complexes is provided in Chapter 20. To oxidize ferrocyanide and ferricyanide most efficiently, ultraviolet light can be employed to photodissociate the complexes, so that lower doses of oxidizing agents will be effective in oxidizing the free cyanide released. This has been demonstrated by Schaefer [18], and the process is examined in depth in Chapter 20. Consider, for example, the photocatalytic dissociation of ferrocyanide and subsequent oxidation of free cyanide and cyanate [18]: − + − Fe(CN)4− 6 + 3H2 O + hυ → 6CN + Fe(OH)3 (s) + 3H + e

(5.36)

CN− + oxidant → CNO−

(5.37)

−

CNO + oxidant →

NO− 3

+ CO2

(5.38)

All of these oxidation reactions follow first-order kinetics [18].

5.3 METAL–CYANIDES: SOLID PHASE COMPOUNDS As outlined in Chapter 2, many solid forms of metal–cyanide species occur, including double metal–metal cyanide complex salts. Solids form upon reaction of free cyanide or metal–cyanide species with metal ions in solution, which results in solid precipitation when the reactants are sufficiently abundant. These solids form and dissolve in natural aquatic systems. Salts of many metal–cyanide species are also synthesized by various routes and sold commercially [62,63]. This section provides detailed information on the physicochemical properties and reactivities of these metal–cyanide species. Solid forms of cyanide can be divided into three forms according to their properties and reactivity. They are (i) simple cyanide solids, like NaCN(s), KCN(s), Ca(CN)2 (s), etc., (ii) alkali or alkaline earth metal–metal cyanide complex solids, like K3 [Fe(CN)6 ](s), Na3 [Fe(CN)6 ] · H2 O(s), etc., and (iii) other metal–metal cyanide complex solids, especially the iron–iron cyanide solids like Fe4 [Fe(CN)6 ]3 (s) and Fe3 [Fe(CN)6 ]2 (s), which are of special interest in water and soil systems.

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5.3.1 SIMPLE CYANIDE SOLIDS These forms of cyanide solids include a metal or ammonium ion bonded ionically to the cyanide anion. These simple cyanide salts generally dissociate readly in aqueous solution to form free cyanide anion, according to the following general reaction: M(CN)x = Mx+ + xCN−

(5.39)

where M represents a metal cation with charge x. Simple cyanides are named because of their structural simplicity and their ability to dissociate substantially under most solution conditions. However, some simple cyanides, where the cyanide ion is bonded to a heavy metal, like CuCN(s) and Zn(CN)2 (s), are somewhat less soluble than others. Most simple cyanide solids release the free cyanide ion readily in aqueous solution. Table 5.5 provides a list of some simple cyanide solids, and some general information on solubility characteristics. Solubility products available in Sehmel [5] for some of the less soluble simple cyanide salts are given in Table 5.6. Only solubility products given in Sehmel [5] are given because these constants have been calculated in a consistent manner and with the most current thermodynamic data (see discussion in Section 5.2.1.1).

5.3.2 ALKALI OR ALKALINE EARTH METAL–METAL CYANIDE COMPLEX SOLIDS These forms of cyanide solids include an alkali (Na+ , K+ ) or alkaline earth (Mg2+ , Ca2+ , Sr 2+ ) metal cation, or ammonium ion (NH+ 4 ), bonded to a metal–cyanide complex. Some of these solids also carry one or more water of hydration within the crystal lattice structure, for example, K4 (FeCN)6 )· 3H2 O(s). These types of complexed cyanide solids are moderately soluble in aqueous solutions under a wide range of pH conditions, releasing metal–cyanide complexes according to the following general reaction: Ax [M(CN)y ] · nH2 O = xA+ + [M(CN)y ]m−

(5.40)

where A represents an alkali or alkaline earth cation or ammonium ion, M is a metal cation complexed with cyanide, x is the number of alkali metal atoms (Equation [5.40] is written for A with charge +1), y is the number of cyanide groups, m is the ionic charge of the liberated metal–cyanide complex, and n is the number of water molecules incorporated in the solid structure. Complex cyanide solids are so named because of their structural complexity and their incomplete dissociation under most solution conditions. The complexed metal–cyanide salts do not dissociate completely, and cyanide is liberated in the form of dissolved metal–cyanide species. Table 5.7 provides a list of some complex cyanide solids, and some general information on solubility characteristics. Solubility products available in Sehmel [5] for some alkali and alkaline earth metal–metal cyanide complex solids are given in Table 5.8. Only solubility products given in Sehmel [5] are given because these constants have been calculated in a consistent manner and with the most current thermodynamic data (see discussion in Section 5.2.1.1).

5.3.3 OTHER METAL–METAL CYANIDE COMPLEX SOLIDS These forms of complex cyanide solids contain one or more transition metals or soft-sphere B-type metal cations [6] bonded to a transition metal–cyanide complex. Some compounds in this class are also referred to as “double metal cyanide complex salts.” The best known such compound is the ferric ferrocyanide solid known as Prussian Blue, Fe4 (Fe(CN)6 )3 (s). Most strong metal–cyanide complexes form stable precipitates under acidic to neutral pH and under excess metal conditions. The iron–iron cyanide solid compounds, presented in Table 5.9, are of special importance in water and soil systems. Where cyanide has been introduced in water and soil systems, these

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TABLE 5.5 Some Simple Cyanide Solids and Solubility Information Aqueous solubilityd (g/100 g H2 O)

Temp (◦ C)

1.7a

15

Formula

Physical forma

Barium cyanide Cadmium cyanide Calcium cyanide Cyanogen bromide Cesium cyanide Cobalt(II) cyanide Cobalt(II) cyanide dihydrate Copper(I) cyanide

Ba(CN)2 Cd(CN)2 Ca(CN)2 CNBr CsCN Co(CN)2 Co(CN)2 · 2H2 O

wh cry pow wh cub cry wh rhom cry; hyg wh hyg needles wh cub cry; hyg blue hyg cry pink-brn needles

CuCN

Copper(II) cyanide Gold(I) cyanide

Cu(CN)2 AuCN

wh pow; or grn orth cry grn pow yel hex cry

Gold(III) cyanide trihydrate Mercury(II) cyanide Nickel(II) cyanide tetrahydrate Potassium cyanide Rubidium cyanide Silver cyanide Sodium cyanide Sodium cyanoborohydride Strontium cyanide tetrahydrate Thallium(I) cyanide Zinc cyanide

Au(CN)3 · 3H2 O

wh hyg cry

Hg(CN)2 Ni(CN)2 · 4H2 O

col tetr cry grn plates

11.4a

25

KCN RbCN AgCN NaCN NaBH3 CN

wh cub cry; hyg wh cub cry wh-gray hex cry wh cub cry; hyg wh hyg pow

69.9a

20

0.0000011 58.2a , 58.7b

20

Sr(CN)2 · 4H2 O

wh hyg cry

TlCN Zn(CN)2

wh hex plates wh pow

0.00047a

20

Solid

Qualitative solubilitya,c vs H2 O; s EtOH s H2 O, EtOH vs H2 O i H2 O i H2 O, acid i H2 O, EtOH; s KCN soln i H2 O; s acid, alk i H2 O, EtOH, eth, dil acid vs H2 O; sl EtOH s EtOH; sl eth i H2 O; s dil acid; s NH4 OH sl EtOH s H2 O; i EtOH, eth i EtOH, dil acid sl EtOH vs H2 O; s, thf; sl EtOH; i bz, eth vs H2 O s H2 O, acid, EtOH reac acid

a Source: Data from Lide, D.R., CRC Handbook of Chemistry and Physics, 85th ed. (online edition), CRC Press, Boca Raton,

FL, 2004. b Source: Data from Dean, J.A., Lange’s Handbook of Chemistry, 15th ed. (online edition), McGraw-Hill, New York, 1999. c Abbreviations (after Lide [82]): ace: acetone; acid: acid solutions; alk: alkaline solutions; aq: aqueous; bz: benzene; col:

colorless; cry: crystals; cub: cubic; dil: dilute; eth: ethyl ether; EtOH: ethanol; grn: green; hex: hexagonal; hyg: hygroscopic; i: insoluble in; orth: orthorhombic; pow: powder; reac: reacts with; rhom: rhombohedral; s: soluble in; sl: slightly soluble in; soln: solution; temp: temperature; tetr: tetragonal; thf: tetrahydrofuran; tol: toluene; vs: very soluble in; wh: white; yel: yellow. d Blanks indicate no values reported in Lide [81] or Dean [82].

compounds often form because of their low solubility and the abundance of iron in the environment [10,40,41,64–67]. Moreover, iron–iron cyanide solids, most notably Prussian Blue, have been produced for use in commercial products such as dyes, inks, pharmaceuticals, and cosmetics for over 150 years [62,63]. As indicated in Table 5.9, various iron–iron cyanide solids have been reported in the literature, with the primary distinguishing feature being the ratio of Fe2+ and Fe3+ in the crystalline structure. The best-known iron–iron cyanide solids are Prussian Blue, Fe4 (Fe(CN)6 )3 (s), and Turnbull’s Blue, Fe3 (Fe(CN)6 )2 (s), which have different proportions of Fe2+ and Fe3+ ; Berlin White or Williamson’s Salt, Fe2 (Fe(CN)6 )(s), which contains all Fe2+ ; and iron–iron cyanide solids containing predominantly Fe3+ , which exhibit colors ranging from brown (Prussian Brown) to green (Berlin Green),

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TABLE 5.6 Equilibrium Solubility Products for Selected Simple Cyanide Solids Solid and dissolution reaction KCN(s) = K+ + CN− NaCN(s) = Na+ + CN− AgCN(s) = Ag+ + CN− AgOCN(s) = Ag+ + OCN− CuCN(s) = Cu+ + CN− Hg(CN)2 (s) + 2H2 O = Hg(OH)02 + 2CN− + H+

Solubility producta (log Ksp at 25◦ C, I = 0) 1.44 2.29 −16.22 −6.62 −19.50 −45.38

a Equilibrium solubility products from Sehmel [5], Table 8.

depending on the relative abundance of Fe3+ and Fe2+ . The intense color of Prussian Blue and similar double metal cyanide complex salts arises due to charge transfer between the metal ions present in different oxidation states [69]. The iron–iron cyanide solids are crystalline coordination compounds with distinct x-ray diffraction spectra [68,69]. Figure 5.15 shows the crystalline structure of Prussian Blue. The unit cell of the dispersed crystalline material is face centered cubic with cyanide ions joining each of the iron ions to its six octahedral nearest-neighbor iron ions [69]. The solubility of the common iron–iron cyanide solids Prussian Blue and Turnbull’s Blue is controlled by the pH and redox potential of the water [41,64–67]. These compounds exhibit low solubility (<1 mg/L) in acidic to neutral pH regimes and under moderately oxic to anoxic conditions. However, under alkaline and highly oxic conditions, these compounds are relatively soluble, yielding corresponding metal–cyanide complexes. The dissolution reactions for Prussian Blue and Turnbull’s Blue are given by: Fe4 (Fe(CN)6 )3 (s) = 4Fe3+ + 3Fe(CN)3− 6

(5.41)

Fe3 (Fe(CN)6 )2 (s) = 3Fe3+ + 2Fe(CN)3− 6

(5.42)

Table 5.10 presents solubility products for various metal–metal cyanide complex solids, including Prussian Blue and Turnbull’s Blue. Note that while the reactions in Table 5.10 are written with products as completely dissociated species for consistency, the metal–metal complex cyanide solids usually do not dissociate completely to produce free cyanide under environmental conditions. The products of dissolution typically are metal–cyanide complexes, as indicated in Equations (5.41) and (5.42). The mass law expressions in Table 5.10 may be combined with appropriate mass law expressions from Tables 5.3 and 5.4 to determine solubility products corresponding to reactions written with metal–cyanide species as dissolution products.

5.4 CYANATE As discussed in Section 5.1.5 and shown in Equations (5.10) and (5.11), free cyanide can be oxidized to form cyanate, CNO− , or, depending on the pH, its protonated form HOCN (pKa = 3.45 at 25◦ C). In CNO− Cyanate and hydrogen cyanate are reported to be substantially less toxic than HCN [2]. Cyanate can also be formed from the hydrolysis of cyanogen halides, such as CNCl and CNBr, which can be formed as disinfection byproducts in waters that are chlorminated [70]; see Chapter 2. Based on the standard redox potentials for the oxidation reactions in Equations (5.10) and (5.11) given in Bard et al. [14], OCN− and HOCN are predicted to be the dominant stable species across

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TABLE 5.7 Some Alkali and Alkaline Earth Metal–Metal Cyanide Complex Solids and Solubility Information Aqueous solubilityd Solid

Formula

Ammonium ferricyanide trihydrate Ammonium ferrocyanide trihydrate Ammonium nitroferricyanide Cobalt(II) ferricyanide Copper(II) ferrocyanide Potassium ferricyanide Potassium ferrocyanide trihydrate Potassium hexacyano cobaltate Potassium silver cyanide Sodium ferricyanide monohydrate Sodium ferrocyanide decahydrate Sodium nitroprusside dihydrate Strontium ferrocyanide pentadecahydrate

Physical forma

(g/100 g H2 O)

Qualitative solubilitya,c

Temp (◦ C)

(NH4 )3 Fe(CN)6 · 3H2 O red cry

s H2 O; i EtOH

(NH4 )4 Fe(CN)6 · 3H2 O yel cry

s H2 O; i EtOH

(NH4 )2 Fe(CN)5 NO

red-brn cry

s H2 O, EtOH

Co3 [Fe(CN)6 ]2

red needles

Cu2 Fe(CN)6 K3 Fe(CN)6

red-brn cub cry or pow red cry

i H2 O, HCl; s NH4 OH i H2 O, acid, os 48.8a , 46b

25a , 20b

K4 Fe(CN)6 · 3H2 O

yel mono cry

36.0a , 28.2b

25a , 20b

K3 Co(CN)6

yel mono cry

vs H2 O; i EtOH

KAg(CN)2

wh cry

s H2 O

Na3 Fe(CN)6 · H2 O

red hyg cry

s H2 O; i EtOH

Na4 Fe(CN)6 · 10H2 O

yel mono cry

i EtOH, eth

20a , 18.8b

20

i os

Na2 Fe(CN)6 NO · 2H2 O red cry

40a

16

sl EtOH

SrFe(CN)6 · 15H2 O

50a

yel mono cry

a Source: Data from Lide, D.R., CRC Handbook of Chemistry and Physics, 85th ed. (online edition), CRC Press, Boca Raton, FL, 2004. b Source: Data from Dean, J.A., Lange’s Handbook of Chemistry, 15th ed. (online edition), McGraw-Hill, New York, 1999. c Abbreviations (after Lide [81]): acid: acid solutions; brn: brown; cry: crystals; cub: cubic; dil: dilute; eth: ethyl ether; EtOH: ethanol; hyg: hygroscopic; i: insoluble in; mono: monoclinic; os: organic solvents; pow: powder; s: soluble in; sl: slightly soluble in; vs: very soluble in; wh: white; yel: yellow. d Blanks indicate no values reported in Lide [81] or Dean [82].

a wide range of pE and pH conditions for the HCN/CN–H2 O system at 25◦ C [2]. Cyanate is rarely encountered in aqueous systems, however, as a strong oxidizing agent and a catalyst are required for conversion of free cyanide to CNO− or HOCN [2]. In addition, cyanate hydrolyzes fairly rapidly, especially under acidic conditions, to carbon dioxide and ammonia [71]. CNO− + 2H2 O → CO2 + NH3 + OH−

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(5.43)

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TABLE 5.8 Equilibrium Solubility Products for Selected Alkali and Alkaline Earth Complex Cyanide Solids Solubility producta (log Ksp at 25◦ C, I = 0)

Solid and dissolution reaction K4 Fe(CN)6 (s) = 4K+ + Fe2+ + 6CN− K4 Fe(CN)6 · 3H2 O(s) = 4K+ + Fe2+ + 6CN− + 3H2 O K3 Fe(CN)6 (s) = 3K+ + Fe3+ + 6CN− K12 Ni8 (Fe(CN)6 )7 (s) = 12K+ + 8Ni2+ + 7Fe2+ + 42CN− K4 Ni4 (Fe(CN)6 )3 (s) = 4K+ + 4Ni2+ + 3Fe2+ + 18CN− K2 Ni3 (Fe(CN)6 )2 (s) = 2K+ + 3Ni2+ + 2Fe2+ + 12CN− K12 Cd8 (Fe(CN)6 )7 (s) = 12K+ + 8Cd2+ + 7Fe2+ + 42CN− K2 CdFe(CN)6 (s) = 2K+ + Cd2+ + Fe2+ + 6CN− K2 Cu2 Fe(CN)6 (s) = 2K+ + 2Cu+ + Fe2+ + 6CN− KZn1.5 Fe(CN)6 (s) = K+ + 1.5Zn2+ + Fe2+ + 6CN− K2 Mn3 (Fe(CN)6 )2 (s) = 2K+ + 3Mn2+ + 2Fe2+ + 12CN− K8 Mn6 (Fe(CN)6 )5 (s) = 8K+ + 6Mn2+ + 5Fe2+ + 30CN−

−48.82 −49.54 −54.64 −431.09 −183.55 −123.13 −441.99 −63.03 −72.51 −66.81 −121.00 −293.68

a Equilibrium solubility products from Sehmel [5], Table 8.

TABLE 5.9 Iron–Iron Cyanide Solids as Reported in the Literature Name

Chemical formula

Color of the solid

Prussian Blue Turnbull’s Blue Prussian Brown Berlin Green Berlin White or Williamson’s Salt

Fe4 [Fe(CN)6 ]3 (s) Fe3 [Fe(CN)6 ]2 (s) Fe[Fe(CN)6 ](s) Not specified Fe2 [Fe(CN)6 ](s)

Bright blue Bright blue Brown Green White

Source: Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Environ. Eng. Sci., 16, 293, 1999. With permission.

The lower the pH, the more rapid the hydrolysis reaction At pH 2, CNO− is hydrolyzed in 5 minutes; at pH 5, 60 minutes; and at pH7 , 22 hours [71]. When cyanate does form it can react with metals, though these reactions have not been studied extensively. Sehmel [5] reports reactions of Ag+ with CNO− to form the dissolved weak metal–cyanide species Ag(OCN)− 2 , and the solid AgOCN(s). Martell et al. [72] report the formation 2− of Co(OCN)4 in addition to the silver complex noted earlier. The reactivity of cyanate has been little studied because it is mostly of interest with respect to its formation in oxidation treatment of free cyanide. Cyanate is readily oxidized to inorganic carbon and nitrogen in the presence of oxidizing agents, as indicated in Equations (5.15) and (5.38), and it hydrolyzes with rate dependent on pH. Cyanate adsorption on soils has not been studied, but it should be retained on soils with significant anion exchange capacity. Based on the adsorption characteristics of similar ions, CNO− is predicted to have some adsorptive affinity for metal oxides in soils (see Figure 5.3).

5.5 THIOCYANATE Free cyanide can react with various forms of sulfur to form thiocyanate, SCN− , which is reported to be relatively nontoxic [2]. The two forms of sulfur in the environment most reactive with CN−

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Fe(III)

N N CC (II)

C

N

C

N

C

N

Fe

N

C

N

N

(III)

Fe

N N CC (II)

Fe

N

C

N

N

(III)

Fe

N N CC (II)

Fe

N

C

N

N

Fe(III)

FIGURE 5.15 Crystal structure for Prussian Blue.

are polysulfides, Sx S2− , and thiosulfate, S2 O2− 3 [2,21]. Typical formation reactions are shown in Equations (5.22) and (5.23). Thiocyanate can protonate to form HCNSo , but this rarely occurs in natural systems as the pKa for this reaction is 1.1 [72]. The occurrence of polysulfides and thiosulfate in water and soils depends strongly on the oxidation–reduction conditions. While thiocyanate can form through reaction of anthropogenic or naturally produced cyanide, most thiocyanate is generated through industrial processes such as coal coking and coal gasification [73], as discussed in Chapters 25 and 26. When present, SCN− is reactive with many metals. Martell et al. [72] report complexes of SCN− with essentially all metals of interest in water and soil. Thiocyanate can be oxidized by oxygen, and when catalyzed by aerobic bacteria the oxidation is rapid. At mid-range pH values, this oxidation reaction may be represented as: + 2− + SCN− + 2O2 + 3H2 O → HCO− 3 + NH4 + SO4 + H

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TABLE 5.10 Equilibrium Solubility Products for Selected Metal–Metal Complex Cyanide Solids Solubility producta (log Ksp at 25◦ C, I = 0)

Solid and dissolution reaction Ag4 Fe(CN)6 · H2 O(s) = 4Ag+ + Fe2+ + 6CN− + H2 O Ag4 Fe(CN)6 (s) = 4K+ + Fe2+ + 6CN− Cd2 Fe(CN)6 · 7H2 O(s) = 2Cd2+ + Fe2+ + 6CN− + 7H2 O Cd2 Fe(CN)6 (s) = 2Cd2+ + Fe2+ + 6CN− Cu2 Fe(CN)6 (s) = 2Cu2+ + Fe2+ + 6CN− Mn2 Fe(CN)6 (s) = 2Mn2+ + Fe2+ + 6CN− Pb2 Fe(CN)6 · 3H2 O(s) = 2Pb2+ + Fe2+ + 6CN− + 3H2 O Pb2 Fe(CN)6 (s) = 2Pb2+ + Fe2+ + 6CN− Tl4 Fe(CN)6 · 2H2 O(s) = 4Tl+ + Fe2+ + 6CN− + 2H2 O Zn2 Fe(CN)6 · 2H2 O(s) = 2Zn2+ + Fe2+ + 6CN− + 2H2 O Zn2 Fe(CN)6 (s) = 2Zn2+ + Fe2+ + 6CN− Fe4 (Fe(CN)6 )3 (s) = 4Fe3+ + 3Fe2+ + 18CN− Fe3 (Fe(CN)6 )2 (s) = 3Fe3+ + 2Fe2+ + 12CN−

−89.69 −193.91 −62.98 −28.22 −61.42 −59.03 −63.60 −27.59 −56.92 −61.23 −29.93 −263.60 −177.40

a Equilibrium solubility products from Sehmel [5], Table 8, except for Fe (Fe(CN) ) (s) and 4 6 3

Fe3 (Fe(CN)6 )2 (s) which are from Ghosh et al. [64,84].

which is the typical reaction that occurs in the oxidation of thiocyanate in nature and in wastewater treatment processes [73]. In SCN− sulfur is in the −1 oxidation state (C is +3, N is −3), whereas sulfur is SO2− 4 is in the +6 oxidation state. Thiocyanate adsorption on soils has been little studied, but it should be retained on soils with significant anion exchange capacity. Based on the adsorption characteristics of similar ions, SCN− is predicted to have some adsorptive affinity for metal oxides in soils (see Figure 5.3).

5.6 ORGANOCYANIDES Organic forms of cyanide are compounds containing the –C≡N group. The most common organocyanides encountered in the environment are the nitriles, R–CN, where R is an alkyl group. The majority of the nitriles are produced synthetically and are widely used as solvents in chemical processing (e.g., acetonitrile, CH3 CN). However, certain classes of nitriles, like cyanohydrins (2-acetoxy-3-butenentrile) and cyanogenic glycosides (amygdalin, dhurrin, and linamarin) occur naturally. Cyanogenic glycosides are found in a wide variety of plant species and take active part in the defense mechanisms of plants (see Chapters 3 and 12 for more details). Certain groups of nitriles, like cyanohydrins, can release free cyanide ions by hydrolysis according to the following reaction [74]: R1

OH

R1 C

C R2

CN

O + CN–

(5.45)

R2

Cyanohydrins are also thermally unstable and decompose to molecular HCN and carbonyl compounds when heated [75]. Cyanogenic glycosides can liberate free cyanide through hydrolysis reactions that frequently are biochemically mediated, usually by enzymatic catalysis. An example

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O C6H10O4 O C6H11O5

O

CH

CH

C6H11O5

CN

CN + C6H12 O6 +H2O

Amygdalin +H2O OH CHO

CH + C6H11 O5 CN

HCN +

Benzaldehyde

Mandelonitrile

FIGURE 5.16 Amygdalin hydrolysis pathway. (Source: Information from: Brimer, L. et al., Riv. Biol., 89, 493, 1996.) glucose

HCN

COO– NH3+

N C

O-glucose

OH

N C

b-glucosidase

OH

OH

Tyrosine

FIGURE 5.17 1552, 1989.)

Dhurrin

O

H

hydroxynitrilelyase

OH

p-hydroxymandelonitrile

OH

p-hydroxy-3 benzaldehyde

Dhurrin catabolism pathway. (Source: Halkier, B.A. and Moller, B.L., Plant Physiol., 90,

of the enzyme-catalyzed hydrolysis of amygdalin is shown in Figure 5.16. It may be seen that the hydrolysis reaction is triggered by the presence of a specific plant enzyme (emulsin) and involves two steps. The sugar moiety first gets cleaved followed by the dissociation of the alkyl cyanide group into aldehyde and cyanide ions [76,77]. Similar to amygdalin, cyanogenesis in dhurrin, a major cyanogenic glycoside in sorghum, is initiated by β-glucosidase which hydrolyzes the cyanogenic glycoside to cyanohydrin (alphahydroxynitrile) and a saccharide. Subsequently the unstable cyanohydrin decomposes spontaneously or enzymatically by the action of a hydroxynitrilelyase to cyanide and a carbonyl compound [78]. This pathway of dhurrin catabolism is depicted in Figure 5.17. In sorghum, the β-glucosidase responsible for hydrolyzing dhurrin is called dhurrinase; this enzyme exists in two isoforms, which show organ-specific expression. The gene (dhr1) encoding dhurrinase-1 has been cloned from sorghum [79], and shows 70% sequence identity to the two maize β-glucosidases. Cyanocobalamin or Vitamin B12 , another common organocyanide compound in which the cyanide ion is directly bonded to a central cobalt cation, can take part in substitution reactions.

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The cyanide ion can be substituted by various anions, forming other forms of Vitamin B12 , namely hydroxycobalamin and chlorocobalamin.

5.7 SUMMARY AND CONCLUSIONS • Free cyanide, consisting of HCN and CN− , is the most reactive (and most toxic) of all the cyanide species. HCN, which dominates free cyanide speciation at pH < 9.2, can exist in both gaseous and aqueous forms, while CN− is present only as the soluble anionic species. The Henry’s Law constant for HCN is relatively high, favoring volatilization. • Weak metal–cyanide complexes (involving Ag, Cd, Cu, Ni, Hg, Zn) are less reactive than free cyanide. These complexes can dissociate under weak acid conditions. They generally occur as anionic species in solution and adsorb most significantly on soils at low- to mid-range pH values. • Strong metal–cyanide complexes (involving Fe, Co, Pt) are less reactive than weak metal–cyanide complexes. The iron–cyanide species ferrocyanide, Fe(CN)4− 6 , and ferri3− cyanide, Fe(CN)6 , are of greatest interest in water and soil systems. These complexes only dissociate under strong acid conditions in the dark, but can dissoicate rapidly upon exposure to light. Ferrocyanide and ferricyanide adsorb most significantly on soils at lowto mid-range pH values. • Metal–cyanide solids form with sufficiently high concentrations of reactants, with formation generally favored at neutral to acidic pH conditions. • Simple cyanide solids like NaCN(s) and KCN(s) are highly soluble in aqueous solutions, releasing the free cyanide anion. • Alkali and alkaline earth metal–metal cyanide complex solids like K4 Fe(CN)6 (s) and K3 Fe(CN)6 (s) are also soluble in aqueous solution, releasing metal–cyanide complexes. • Double metal–metal cyanide complex solids, like Prussian Blue, Fe4 (Fe(CN)6 )3 (s), and Turnbull’s Blue, Fe3 (Fe(CN)6 )2 (s), are usually less soluble under acidic to neutral pH conditions. However, under alkaline pH conditions, these complex solids become more soluble, releasing metal–cyanide complexes. • Cyanate, CNO− , can be formed from oxidation of free cyanide. It rarely occurs in natural water and soil systems, however, as a strong oxidizing agent and a catalyst are required for the oxidation of free cyanide to proceed at an appreciable rate. Also, CNO− hydrolyzes rapidly. • Thiocyanate, SCN− , can be formed from reaction of free cyanide with polysulfides, Sx S2− , and thiosulfate, S2 O2− 3 . While thiocyanate can form through reaction of anthropogenic or naturally produced cyanide, most thiocyanate is generated through industrial processes. When present in soil and water systems, SCN− can form complexes with many metals. • Organocyanide compounds are usually stable in nature. Organocyanide compounds can undergo hydrolysis reactions releasing free cyanide. Certain organocyanides, like, cyanocobalamin can also undergo substitution reactions, where the cyanide moiety is replaced by other inorganic anions.

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79. Cicek, M. and Esen, A., Structure and expression of a dhurrinase (B-glucosidase) from sorghum, Plant. Physiol., 116, 1469, 1998. 80. Doudoroff, P., Leduc, G., and Schneider, C.R., Acute toxicity to fish of solutions containing complex metal cyanides, in relation to concentrations of molecular hydrocyanic acid, Trans. Am. Fish. Soc., 95, 6, 1966. 81. Avedesian, M.M., Spira, P., and Kanduth, H., Stripping of HCN in a packed tower, Can. J. Chem. Eng., 61, 801, 1983. 82. Lide, D.R., CRC Handbook of Chemistry and Physics, 85th ed. (online edition), CRC Press, Boca Raton, FL, 2004. 83. Dean, J.A., Lange’s Handbook of Chemistry, 15th ed. (online edition), McGraw-Hill, New York, 1999. 84. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Clarification: equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 501, 1999. 85. Halkier, B.A. and Moller, B.L., Biosynthesis of cyanogenic glucoside dhurrin in seedlings of Sorghum bicolor (L.) (Moench) and partial purification of the enzyme system involved, Plant Physiol., 90, 1552, 1989.

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Transformation of 6 Biological Cyanide in Water and Soil Stephen D. Ebbs, George M. Wong-Chong, Brice S. Bond, Joseph T. Bushey, and Edward F. Neuhauser CONTENTS 6.1

6.2

6.3

6.4

6.5

6.6

Microbial Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.1 Bacteria Capable of Degrading Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.2 Conditions Required for Bacterial Degradation of Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.3 Pathways for Bacterial Degradation of Cyanogenic Compounds . . . . . . . . . . . . . . . 6.1.3.1 Bacterial Degradation of Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.3.2 Bacterial Degradation of Nitriles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.3.3 Bacterial Degradation of Metal–Cyanide Complexes . . . . . . . . . . . . . . . 6.1.3.4 Bacterial Degradation of Thiocyanate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fungal Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.1 Fungi Capable of Degrading Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2.2 Conditions Required for Fungal Degradation of Cyanogenic Compounds. . . . . Pathways for Fungal Degradation of Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.1 Fungal Degradation of Free Cyanide and Nitriles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.2 Fungal Degradation of Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3.3 Fungal Degradation of Thiocyanate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bacterial Assimilation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.1 Bacteria Capable of Assimilating Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . 6.4.2 Conditions Required for Bacterial Assimilation of Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.3 Pathways for the Bacterial Assimilation of Cyanogenic Compounds . . . . . . . . . . 6.4.3.1 Bacterial Assimilation via the β-cyanoalanine Synthase Pathway. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.3.2 Bacterial Assimilation via γ-Cyano-α-aminobutyric Acid Synthase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fungal Assimilation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.1 Fungi Capable of Degrading Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.2 Conditions Required for the Fungal Assimilation of Cyanogenic Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5.3 Pathways for the Fungal Assimilation of Cyanogenic Compounds . . . . . . . . . . . . 6.5.3.1 Fungal Assimilation via α-Aminobutyric Acid Synthesis . . . . . . . . . . 6.5.3.2 Alanine and Glutamate Synthesis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Botanical Transformation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.1 Cyanide-Resistant Respiration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.2 Plant Pathways for the Assimilation of Cyanogenic Compounds . . . . . . . . . . . . . .

94 95 95 96 96 98 98 100 101 101 101 101 101 102 103 104 104 104 105 105 106 106 106 107 107 107 108 108 109 110 93

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6.6.2.1 6.6.2.2 6.6.2.3 6.6.2.4 6.6.2.5

Plant Assimilation via the Cyanoalanine Pathway . . . . . . . . . . . . . . . . . . Plant Assimilation via the Sulfur Transferase Pathway . . . . . . . . . . . . . Plant Assimilation via Cyanase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Plant Assimilation of Cyanamide Herbicides . . . . . . . . . . . . . . . . . . . . . . . Plant Transport and Transformation of Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.2.6 Volatilization of Cyanide by Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.7 Environmental Impacts of Cyanide Assimilation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.8 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

110 111 111 112 113 114 114 115 115

Cyanide has long held a prominent position amongst metabolic inhibitors, best known as an inhibitor of aerobic respiration [1]. However, given the ubiquity of cyanide and cyanogenic (cyanide-containing) compounds in nature, it is not surprising that these compounds can be biologically transformed or utilized. The metabolic capacity to transform cyanogenic compounds no doubt stems from the substantial impact that cyanide had on the early development of life in a Hadean atmosphere, which contained a significant concentration of cyanide. These metabolic pathways for cyanide transformation are not only influenced by the presence of cyanide but also by environmental factors, including pH, available carbon sources, and oxygen. This chapter presents a discussion of the microbial, fungal, and botanical pathways for the transformation of cyanide and cyanogenic compounds. Biological transformation in the context of this presentation is discussed within two contexts — degradation and assimilation. Degradation refers here to reactions that convert cyanide into simple organic or inorganic molecules, which can be further metabolized to ammonia and either carbon dioxide or methane. Degradation pathways are primarily utilized by prokaryotic organisms and likely represent one of the earliest evolutionary responses to the presence of cyanide in the environment. Assimilation, in the context of this chapter, refers to the incorporation of the cyanide moiety directly into organic compounds. Assimilatory pathways are present in both prokaryotes and eukaryotes, but primarily represent the pathways for cyanide transformation in the latter. This movement toward assimilation in eukaryotes may signal a metabolic change from cyanide detoxification to a more energy-efficient utilization of cyanogenic compounds as a source of both carbon and nitrogen. Regardless, the transformation of cyanide compounds is a natural biological process, which can be accomplished through a diverse array of biochemical reactions mediated by numerous species of micro-organisms. This diversity of reactions has also led to the development of a variety of technologies for the biological processing of cyanide contaminated wastes and groundwater as discussed in Chapters 23, 24, 26, and 27.

6.1 MICROBIAL DEGRADATION Microbial degradation/utilization of cyanide and cyanogenic compounds has been investigated since the mid-1950s [2,3]. Since that time, much has been learned about the micro-organisms responsible for cyanide degradation, the enzymatic pathways involved, and the conditions under which the degradation of cyanogenic compounds occurs. The degradation of cyanide and cyanogenic compounds is generally induced by the presence of the specific substrate (e.g., free cyanide, thiocyanate) and therefore represents an effort to detoxify cyanide, and ultimately to convert the carbon and nitrogen present into a form that can be utilized by the bacteria. Several reviews of microbial degradation of cyanide can be found in the literature [4–7].

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TABLE 6.1 A Partial List of Bacteria and Fungi Capable of Metabolizing Free Cyanide and Cyanogenic Compounds and the Substrate from which the Strains were Isolated Bacteria Soil Alcaligenes spp. Arthrobacter spp. Bacillus pumilus Burkholderia cepacia Chromobacterium violaceum Citrobacter freundii Fungi Soil Fusarium oxysporum N-10 Candida guilliermondii Fusarium solani Gloeocercospora sorghi Leptosphaeria maculans Penicillum miczynski

Pseudomonas aeruginosa Pseudomonas fluorescens Pseudomonas putida Thiobacillus thiocyanoxidans Thiobacillus thioparus

Sewage sludge Bacillus spp. Norcardia rhodochrous

Waste streams Bacillus pumilus Escherichia coli Klebsiella oxytoca Pseudomonas nonliqurfaciens Pseudomonas stutzeri

Sewage sludge

Waste streams Acremonium strictum Cryptococus humicolus MCN2 Fusarium oxysporum

Rhodococcus sp. Rhizopus oryzae Scytalidium thermophilum Stemphylium loti Trichoderma polysporum

Organisms isolated from soil include those that are phytopathogenic, were isolated from plant wastes, or were identified in contaminated soils or sediments. Waste streams refers primarily to industrial effluents.

6.1.1 BACTERIA CAPABLE OF DEGRADING CYANOGENIC COMPOUNDS Table 6.1 lists bacterial species that have been identified as having the capacity to have appropriate enzyme systems induced for degrading cyanide and cyanogenic compounds. Past study has shown that some strains of these species (e.g., Chromobacterium violaceum, Bacillus megaterium, and Escherichia coli) possess more than one pathway for the degradation of cyanide compounds. The table illustrates that bacteria capable of degrading cyanogenic compounds are found in the natural environment as well as contaminated soil, water, and wastes. Their presence in contaminated media, though, has already provided a selection pressure that enriches cyanide-degrading organisms in the microbial consortia, facilitating their subsequent isolation. Most of the progress in the study of the microbial degradation of cyanogenic compounds has come from such studies.

6.1.2 CONDITIONS REQUIRED FOR BACTERIAL DEGRADATION OF CYANOGENIC COMPOUNDS The degradation of cyanogenic compounds occurs under a wide variety of physicochemical conditions. While there is ample evidence of degradation under near neutral pH values, degradation of cyanide [8] and thiocyanate [9] under highly alkaline conditions (pH >9.0) and metal–cyanide compounds under acidic (pH 4.0) conditions [10,11] has also been observed. Oxygen is not absolutely required as degradation can occur under anaerobic conditions [12–15]. Degradation of cyanogenic compounds can occur in the presence of a variety of carbon sources, including simple sugars such as glucose [8], acetate [9], sugarcane molasses [16], and phenol [15,17,18]. Cyanide can in many instances be utilized as the sole source of nitrogen by some microbial strains, as free cyanide [19], nitriles [20], metal–cyanide complexes [21], or thiocyanate [22]. This is supported by the observation that the degradation of cyanogenic compounds generally coincides with log-phase growth

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of bacterial cultures [23]. The extent to which cyanogenic compounds sustain microbial growth in natural systems has not been fully characterized, but the laboratory data suggests that cyanogenic compounds could fulfill this role if the concentration in the media remained high. As cyanogenesis (biological production of cyanide compounds) is as ubiquitous in nature as cyanide transformation (see Chapters 3 and 12), organisms capable of degrading cyanogenic compounds may rely upon cyanogenic organisms for these compounds. This relationship between cyanogenesis and cyanide transformation has produced examples of cyanide “microcycles” in nature, which are discussed in greater detail in Section 6.7.

6.1.3 PATHWAYS FOR BACTERIAL DEGRADATION OF CYANOGENIC COMPOUNDS Cyanogenic compounds can provide both nitrogen and carbon for microbial growth. In some instances, the cyanide compound can be the sole source of carbon and nitrogen supporting growth. Nevertheless, the pathways utilized by bacterial species to degrade various compounds (e.g., free cyanide, metallocomplexed cyanide, thiocyanate, and organonitriles) differ and may depend on other physicochemical characteristics of the medium. Most of the information gathered on microbial degradation of cyanide compounds was obtained from in vitro laboratory experiments (e.g., shaker flask tests; batch-fed or continuous flow reactors) with cultures isolated by enrichment procedures. Culture media were generally standard mineral or broth type at pH near neutral and incubation at 25 to 30◦ C. The extent to which cyanide degradation occurs under natural conditions has not been explored in great detail.

6.1.3.1 Bacterial Degradation of Free Cyanide There are four general pathways used by micro-organisms to transform free cyanide — oxidative, reductive, hydrolytic, and substitution/transfer (Table 6.2). The last is an assimilatory pathway discussed in Section 6.4. The first three pathways are degradative pathways in that they convert cyanide into simpler molecules (e.g., formamide, formic acid, and ammonia) that can then be assimilated, with the expenditure of metabolic energy by the micro-organism, as sources of carbon and nitrogen for growth. In some species, organisms, or strains, more than one pathway may be utilized for cyanide biodegradation or assimilation [6]. The activity of the pathway(s) is dictated by external conditions (i.e., aerobic and anaerobic conditions). For example, there is evidence that cyanide biodegradation under anaerobic conditions may represent a secondary reaction or cometabolism, depending on the form of reduced carbon present in the system and the microbial consortium [6,15]. As for other organic contaminants, cyanide bioavailability and solubility in soil–water systems is a determining factor for cyanide biodegradation [24]. Finally, some reactions require cofactors or reducing equivalents. The oxidative reactions require NAD(P)H to catalyze the oxygen-dependent biodegradation of cyanide in the presence of a carbon source. Two possible pathways have been proposed for this reaction. In the first, a cyanide monoxygenase converts cyanide to cyanate (Table 6.2). A cyanase enzyme then catalyzes the bicarbonate-dependent conversion of cyanate to ammonia and carbon dioxide. The enzymatic oxidation of free cyanide is stoichiometric with one mole of cyanide producing one mole each of carbon dioxide and ammonia [25–28]. Each reaction is inducible only by its respective substrate. Although cyanide monoxygenase has been identified in a limited number of species (e.g., Pseudomonas fluorescens [25]), cyanases have been identified in numerous species of bacteria and plants, with related DNA sequences in fungi and animals [29]. The presumed role of cyanase has long been as a protection against cyanate poisoning [6]. However, as cyanate is not a common metabolite, there has been speculation that cyanases may have additional fundamental roles in bicarbonate/CO2 metabolism and nitrogen metabolism [29].

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TABLE 6.2 The General Categories of Chemical Reactions Responsible for the Biodegradation of Free Cyanide and Thiocyanate Free Cyanide and nitriles Hydrolytic reactions Cyanide hydratase HCN + H2 O → HCONH2

Cyanidase HCN + 2H2 O → HCOOH + NH3

Nitrile hydratase R–CN + H2 O → R–CONH2

Nitrilase R–CN + 2H2 O → R-COOH + NH3

Oxidative reactions Cyanide monoxygenase HCN + O2 + H+ + NAD(P)H → HOCN + NAD(P)+ + H2 O Cyanide dioxygenase HCN + O2 + 2H+ + NAD(P)H → CO2 + NH3 + NAD(P)+ Reductive reactions HCN + 2H+ + 2e− → CH2 =NH → CH2 =NH + H2 O → CH2 =O ↓ CH2 =NH + 2H+ + 2e− → CH3 –NH + 2H+ + 2e− → CH4 + NH3 Thiocyanate Carbonyl pathway (thiocyanate hydrolase) SCN− + 2H2 O → COS + NH3 + OH− Cyanate pathway (cyanase) + SCN− + 3H2 O + 2O2 → CNO− + HS− → HS− + 2O2 → SO2− 4 +H ↓ + CNO− + 3H+ + HCO− 3 → NH4 + 2CO2 For the hydrolytic reaction involving nitriles, R– represents either an aliphatic or aromatic group. The cyanate formed by cyanide monoxygenase is converted to NH+ 4 and CO2 by the same pathway as the cyanate from thiocyanate. The reductive pathway is derived from the action of nitrogenase and the products resulting from the transfer of pairs of electrons.

A second oxygenase pathway observed in Pseudomonas spp., E. coli, and Bacillus pumilius involves a single enzymatic reaction catalyzed by cyanide dioxygenase (Table 6.2). This reaction also requires NAD(P)H, a pterin (nitrogen heterocyclic ring related to the nucleic acid guanidine) cofactor [30], and oxygen, but forms ammonia and carbon dioxide directly in a single step reaction. Studies of E. coli strain BCN6 and P. fluorescens NCIMB 11764 suggested that the formation of cyanohydrin (glucose–cyanide) complexes was necessary for cyanide biodegradation via this oxygenase pathway [31,32], but the need for this complexation has not been definitively demonstrated. The authors suggested that the formation of these complexes may represent a more generalized pathway utilized by organisms for cyanide tolerance and utilization [31]. The study of these oxygenase pathways has been facilitated by the identification of a P. fluorescens strain (JL102) defective in cyanide oxygenase activity [33]. Curiously, study of this mutant strain has also suggested that a nonenzymatic siderophore-based mechanism may also contribute to cyanide utilization [19]. The reductive pathway for the degradation of free cyanide is generally considered to occur under anaerobic conditions, although this is perhaps not universally true. For example, there is a report of hydrolytic degradation of free cyanide under anaerobic conditions by some bacteria [34,35]. Current information suggests that the reductive pathway is mediated by the bacterial nitrogenase enzyme [36,37] in species such as Rhodopseudomonas gelatinosa and Klebsiella oxytoca. Nitrogenase, the enzyme used by micro-organisms during nitrogen fixation accepts a variety of substrates, including HCN. When HCN is utilized by nitrogenase, the subsequent reactions lead to the

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production of ammonia, as for the oxidative pathways, and methane. Methane production can be high under anaerobic batch conditions, and has prompted some investigators to explore the possibile utility of reductive cyanide degradation as a technology for biogas production from cyanide containing wastes [12,15,38]. Hydrolytic cyanide biodegradation reactions are catalyzed by either cyanide hydratase or cyanidase. The difference between the two hydrolytic pathways is the product formed (Table 6.2). The reaction mediated by cyanide hydratase forms formamide while cyanidase forms formate and ammonia. These two enzymes utilize HCN as their substrate, rather than the cyanide anion [39]. Cyanide hydratase is primarily a fungal enzyme, although it has been identified in a strain of P. fluorescens [35]. In contrast, cyanidases (also referred to as cyanide dihydratases) appear principally in bacteria. Cyanidase reactions are believed to form formate directly. Although these strains do not metabolize formamide when provided in the substrate, an enzyme-bound formamide intermediate in the cyanidase reaction has not been excluded [6]. Pseudomonas fluorescens NCIMB 11764 is unique in that this strain has been shown to biodegrade cyanide (added as KCN) by at least three different pathways [35], depending upon cyanide concentration and oxygen availability. Activity consistent with both cyanidase and cyanide hydratase activity were observed under aerobic and anaerobic conditions as production of formamide and formate. Subsequent metabolism of formamide was not observed in culture, so formamide production was considered to be the terminal product of a parallel pathway. 6.1.3.2 Bacterial Degradation of Nitriles Cyanide hydratase and cyanidase have a significant structural similarity at the amino acid and protein level to nitrilase and nitrile hydratase enzymes [39]. Like the corresponding enzymes for cyanide, nitrilases and nitrile hydratases convert organic nitriles (e.g., R–CN) to the corresponding acid or amide, respectively (Table 6.2). While the enzymes for HCN degradation are specific for this substrate, the enzymes capable of degrading nitriles show less substrate specificity. Nitrile hydratases can effectively degrade a wide variety of structurally diverse nitriles, primarily aliphatic nitriles [40] that are neutral or positive in charge and not highly substituted [41]. The ability to degrade aromatic nitriles via nitrile hydratase appears limited to one species, Nocardia rhodochrous LL100-21 [42]. A general hydratase/amidase system was identified in Rhodococcus sp. CH5 and displayed a capacity to transform >20 organic nitriles, including propionitrile, acrylonitrile, benzonitrile, succinonitrile, methyl cyanoacetate, and others [41]. For the hydratase/amidase system, the nitrile hydratase converts the nitrile to an amide intermediate, which is then converted to the corresponding acid and ammonia by an amidase. Similar reaction pathways have been observed in species of Escherichia, Pseudomonas, Brevibacterium, and Acinetobacter [41,43,44]. In contrast, nitrilase enzymes such as those in Pseudomonas and Nocardia are specific to aromatic and heterocyclic nitriles [40,41] and are unable to hydrolyze aliphatic nitriles. 6.1.3.3 Bacterial Degradation of Metal–Cyanide Complexes The microbial degradation of metal–cyanide complexes has long been reported in the literature. This subject has garnered some skepticism as few studies have provided adequate controls for lightmediated dissociation (photodissociation) or physical dissociation of the metal complexes [45,46]. Many also do not necessarily relate growth in media containing metal–cyanide complexes specifically to the capacity to degrade those complexes [43,44], which does not account for the possible utilization of cyanide liberated from dissociation of the complex. The discussion here focuses upon the reported examples of metal–cyanide degradation, regardless of the degree of control provided for these two factors. The susceptibility of metal–cyanide complexes to microbial degradation reportedly parallels the chemical stability of the complex itself [2]. Cyanide complexes with cadmium, copper, nickel, silver, and zinc show much less stability than gold and iron complexes (see Chapter 5). Similarly,

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studies have shown that cyanide complexes with iron and gold are more resistant to microbial degradation [47,48]. Cyanide complexes with other metals apparently support microbial growth, although the degradation rates differ depending on the individual metal complex. Studies have reported that zinc and cadmium complexes were totally degraded; copper and nickel complexes were 75 to 80% degraded, and ferrocyanide was only slightly degraded [2]. Concentrations of the metal–cyanide complex in excess of 100 mg l−1 , as well as high concentrations of the metal cation, inhibit microbial degradation [2,46], as does the presence of ammonia [49]. Cyanide complexes with nickel (e.g., tetracyanonickelate(II), or TCN) were reportedly degraded by strains of Pseudomonas and Klebsiella isolated from enrichment cultures obtained from soils with no previous history of cyanide contamination [49,50]. The strains examined, including P. putida BCN3, could utilize this metal cyanide complex as the sole source of nitrogen for growth, with increases in growth occurring in parallel with loss of TCN from the media [49]. Interestingly, degradation of TCN was accompanied by the formation of Ni(CN)2 , as verified via FT-IR spectroscopy. No subsequent degradation of the Ni(CN)2 was reported. The authors were unable to determine if TCN served as a direct substrate or if dissociation to yield CN− was required to provide nitrogen for growth, although the latter was noted as being the more likely explanation. TCN reportedly induced cyanide oxygenase activity [50], but there was no definitive evidence to determine whether this was due to specifically the presence of TCN or induced following the dissociation of CN from TCN. Several studies reported biodegradation of cyanide complexes with copper and zinc [43,51,52]. Species of Enterobacter, Pseudomonas [51], and Escherichia [43] were shown to grow on media containing these complexes as a nitrogen source. For the Enterobacter and Pseudomonas species, this growth was accomplished in batch or continuous-fed reactors in a mining wastewater, which also contained free cyanide, thiocyanate, and various weak-acid dissociable cyanide species. Degradation was assumed based upon the apparent loss of metal–cyanide compounds from solution and the generation of ammonia within the reactor. Growth of E. coli BC6 was achieved with copper cyanide as the sole nitrogen source, although concentrations of <5 mg l−1 restricted bacterial growth. This same strain was able to grow on zinc cyanide concentrations of up to 100 mg l−1 , but with decreases in growth observed at concentrations >10 mg l−1 [43]. Control media, containing metal–cyanide complexes in the absence of a bacteria inoculum showed losses of total cyanide as high as 60%. In the presence of this E. coli strain, there was 100% loss of both copper and zinc complexes. A bacterial consortium, comprised of a 17:1:1:1 mixture of four bacterial species (Citrobacter specie MCM B-181 and Pseudomonas species MCM B-182, MCM B-183, and MCM B-184) purportedly degraded nearly 100% of the cyanide complexes with copper, silver, or zinc added to a continuous flow rotating biological contactor in 15 h [16,52]. The pH optimum for this degradation was ∼7.5. The cyanide complexes provided the sole nitrogen source with molasses as the carbon source. As in other studies, degradation of the metal–cyanides was inhibited by 15 to 50% when ammonia was added to the system. Iron cyanide complexes are among the most stable of metallo-cyanide complexes. While lightsensitive, these compounds undergo low levels of dissociation in the dark; this dissociation is influenced by pH and other water quality factors [53]. Nevertheless, there are studies, which suggest that these compounds can be degraded by bacteria, regardless of whether the strains have experienced a previous exposure to cyanide. In one instance, a mixed consortium isolated from the soil of a former manufactured gas plant site and with a demonstrated capability to degrading free cyanide was grown in aqueous solution and soil slurries in the presence of potassium hexacyanoferrates(II) [24]. In the absence of light at a pH of 7.2, there was a ≥60% loss of free cyanide from the inoculated solution systems in 357 h. In contrast, there was only a 7 to 15% reduction in free cyanide concentration in the soil slurries, demonstrating the importance of substrate bioavailability to degradation. Chemical dissociation of the complex followed by microbial degradation of free cyanide could not be discounted in this study. Bacterial cultures enriched from soil without cyanide or from standardized ATCC cultures also reportedly degraded iron cyanide complexes. Specifically, degradation of ferrocyanide by

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P. fluorescens P70 was examined in an aerobic batch fermenter [46] while potassium hexacyanoferrate degradation by Pseudomonas aeruginosa (ATCC 27883) or E. coli (ATCC 259277) was studied in simple glass reactors painted black to exclude light [45]. Degradation by P. fluorescens P70 was optimal at pH 5.0 and required the presence of glucose as a carbon and energy source. Ferrocyanide purportedly provided the sole nitrogen source for growth, even when present at only 100 mg l−1 (as ferrous(II) cyanide complex) in solution. In a solution of sterilized river water or commercial nutrient broth containing 3293 mg l−1 potassium hexacyanoferrate(III) and inoculated with P. aeruginosa and E. coli, free cyanide concentrations increased sharply after four days to a concentration >1.5 mg l−1 , followed by gradual decreases for the following 21 days. This increase in free cyanide concentration was accompanied in both systems by a 2–3 unit decrease in pH, reaching values of 5.0–5.5 in three to four days. An investigation of the ability of proprietary aerobic bacterial strains to degrade iron–cyanide complexes was performed with laboratory batch reactors and solutions of potassium ferrocyanide, spent potliner leachate, and groundwater affected by spent potliner leachate [47]. The bacterial strains were added to the batch reactors, in which environmental conditions (oxygen and nutrient levels, pH, temperature) were carefully controlled. For test periods up to 200 h, there was no indication of degradation of potassium ferrocyanide in the synthetic test solution, or of total cyanide in the spent potliner leachate or groundwater. Monitoring of the bacterial populations by enumeration showed that the bacteria were tolerant of the cyanide species present at concentrations from 10 to 1000 mg l−1 , but there was no evidence that the strains could grow on these substrates as a sole source of carbon and energy. While there were clear losses of cyanide from the systems utilized in these studies, it was not possible to demonstrate conclusively that the micro-organisms were directly responsible for the degradation of the hexacyanoferrates. In these studies, parallel efforts with sterile systems generally showed that neither time nor pH contributed substantially to the increase in free cyanide concentration. Nevertheless, free cyanide appears in the media in the presence of bacteria. The mechanism by which this occurs remains a mystery. Unambiguous evidence for direct degradation of hexacyanoferrates, or other metal–cyanide complexes for that matter, is lacking. An alternate explanation may be that bacterial exudates promote dissociation of the complexes and that these bacterial strains degrade the liberated free cyanide. While clarifying this specific question will require additional study, this does not necessarily preclude the use of bacteria in biotreatment systems for metal–cyanide complexes. For example, the treatment system at the Homestake Lead, ND consistently remove 91 to 95% total cyanide (which includes iron and other metal complexes); significant portion of the iron complex was removed by adsorption onto the microbial mass [54]. The wastewater processed also contained significant concentration of thiocyanate [6]. 6.1.3.4 Bacterial Degradation of Thiocyanate Thiocyanate is typically considered a product of the biological transformation of cyanide (Section 6.6.2.2), rather than as a substrate for degradation. Nonetheless, bacterial degradation of thiocyanate has been reported and two possible pathways have been proposed (Table 6.2). Most evidence supports a cyanate pathway, in which thiocyanate is converted first to cyanate, with subsequent formation of ammonia and carbon dioxide. Evidence for this pathway has been obtained from studies of several bacterial species [9,51,55–59], including Thiobacillus thiocyanoxidans and Pseudomonas stutzeri. Most of these studies demonstrated that thiocyanate was used as source of carbon, nitrogen and, in some instances, sulfur. Thiocyanate degradation occurs under a range of physicochemical conditions, including highly alkaline soda lake sediments and soils [9]. A carbonyl pathway has also been reported. In this pathway, thiocyanate is converted directly to ammonia and carbonyl sulfide. This pathway was observed in the chemolithotroph Thiobacillus thioparus. Further efforts allowed for the identification of the enzyme responsible (thiocyanate hydrolase) and genes encoding this enzyme in other thiocyanate-degrading bacterial cultures [60]. Interestingly, this

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hydrolase shows a significant homology to nitrile hydratases, suggesting some similarity between nitrile and thiocyanate degradation.

6.2 FUNGAL DEGRADATION 6.2.1 FUNGI CAPABLE OF DEGRADING CYANOGENIC COMPOUNDS Much of the evidence for fungal degradation of cyanide was obtained from the study of pathenogenic fungi, particularly those that attack cyanogenic plants. The capacity to degrade cyanide evolved in these species, including Stemphylium loti, Gloeocercospora sorghi, Leptosphaeria maculans, and Helminthosporium trucicum, as a means of countering the cyanogenic chemical defenses of plants. The fact that fungal degradation of cyanide is more prevalent in pathogenic fungi than nonpathenogenic [61] supports this contention. Nevertheless, the capacity to degrade cyanogenic compounds has been observed in other fungi as well, including species from the genera Fusarium [62,63], Trichoderma [64], Cryptococcus [21,65], Scytalidium, and Penicillium [11]. There is also a report of cyanide degradation by white rot fungi [66], attributed to the activity of lignin-degrading enzymes. Since this mechanism differs greatly from those identified in the aforementioned fungi and there has been no further support for this pathway, it is not discussed here.

6.2.2 CONDITIONS REQUIRED FOR FUNGAL DEGRADATION OF CYANOGENIC COMPOUNDS Fungal degradation occurs under conditions generally similar to those for bacteria. Degradation is commonly observed at neutral pH, although there are reports of cyanide degradation at pH 4.0 [11] and under alkaline conditions up to pH 10.7 [6,62]. The extent of degradation is lower under alkaline conditions than at neutral pH, most likely due to changes in cyanide speciation. Like the hydrolytic pathways of bacteria, the hydrolytic fungal degradation pathway preferentially utilizes HCN over CN− as a substrate [39]. Cyanide speciation shifts from HCN to CN− as the pH increases, suggesting that the decrease in degradation may be due to substrate availability. Fungal degradation is enhanced by the presence of HCN. The presence of simple sugars and other organic nutrients in the medium enhances fungal degradation, while little to no growth is observed in media without these compounds. Apparently cyanide alone (as the sole source of carbon and nitrogen) cannot support fungal growth [62]. Unlike the bacterial oxygenase pathways, fungi require no addition cofactors as growth and degradation on basal complete media containing cyanide was possible.

6.3 PATHWAYS FOR FUNGAL DEGRADATION OF CYANOGENIC COMPOUNDS The fungal pathways for the degradation of cyanogenic pathways are more limited than those in bacteria and in some instances show mechanistic differences. While fungi show the capacity to degrade free cyanide, nitriles, and perhaps metal–cyanide complexes, this capacity is reportedly limited to a single hydrolytic pathway. Thiocyanate degradation has also been observed and involves the cyanate pathway described earlier for bacteria.

6.3.1 FUNGAL DEGRADATION OF FREE CYANIDE AND NITRILES Fungal degradation of free cyanide is a two-step hydrolytic process, in which cyanide hydratase (also called formamide hydrolase) converts HCN to formamide. An amidase such as formamidase is then responsible for the formation of formic acid and ammonia (Table 6.2). Both steps of the pathway are obligate for cyanide degradation as fungi such as Fusarium solani are unable to grow on either

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formamide or formic acid as the sole sources of carbon and nitrogen [62]. In fact, growth in media supplemented with cyanide and nutrients does not occur until the formamide has been converted to formic acid. Activity of the first enzyme, cyanide hydratase, is induced only in the presence of cyanide [63]. The reaction is irreversible with a pH optimum in the 7.0 to 9.0 range [5,62]. Cyanide hydratase purportedly requires no energetic cofactor and shows little dependence upon nutrients from the media. The cyanide hydratases from F. lateritium and G. sorghi show 65% homology at the gene sequence level and 75% homology at the amino acid level, indicating a high degree of similarity between fungal enzymes. The fungal enzyme shows only 35% homology to the cyanide hydratase from Alcaligenes faecalis [63]. The second reaction is dependent upon nutrients in the media as well as oxygen. When F. solani was grown in minimal media in the presence of cyanide, formamide is the sole product formed. However, when yeast extract was present the formamide was converted to formic acid after a lag period [62]. The rate of the amidase reaction is slower than that of cyanide hydratase. For example, in F. solani the formation of formic acid from formamide was only 25% of the rate of formamide formation from HCN. The identity of this amidase has not been conclusively demonstrated, although it has been suggested to be a formamidase similar to those from A. xylosoxidans and P. putida. Fungal degradation of nitriles has also been demonstrated. Two yeast strains, Cryptococcus sp. UFMG-Y28 and Candida guilliermondii CCT 7207, were able to use nitriles such as benzonitrile as the sole nitrogen source for growth [65]. Both free and immobilized cultures of Candida guilliermondii CCT 7207 degraded aliphatic and aromatic nitriles, although immobilized cultures showed a slightly slower growth and acetic acid formation rate [67]. However, immobilization allowed for degradation under a higher initial nitrile concentration.

6.3.2 FUNGAL DEGRADATION OF METAL–CYANIDE COMPLEXES Evidence for fungal degradation of metal–cyanide complexes is comparable to that for bacteria. That is, there are studies that demonstrate growth of fungal cultures in media containing metal cyanides as the nitrogen source concomitant with losses of cyanide from solution. A mixed fungal culture consisting of Fusarium solani, Trichodera polysporum, Fusarium oxysporum, Scytalidium thermophilum, and Penicillum miczynski isolated from a former gasworks site was capable of using tetracyanonickelate (TCN) and hexacyanoferrate as nitrogen sources at pH 4.0 and 7.0 but not at alkaline pH [10,11]. Similar results were obtained for Fusarium oxysporum N-10 [68] and Cryptococcous humicolus MCNZ [21] when grown on TCN. The results of these studies have suggested a role for cyanide hydratase in the fungal degradation of metal cyanides. In most of these studies, formic acid and ammonia were reported as terminal degradation products, with a formamide intermediate detected in some instances. For the transformation of TCN by F. solani and T. polysporum at pH 7, the rates of ammonia evolution and the rates of TCN dissociation in solution were almost identical (∼1.9 mmol min−1 l−1 ) [10]. Carbon dioxide, most likely arising from the metabolism of formic acid, has also been observed [10], a result consistent with the cyanide hydratase pathway. Degradation of both tetracyanonickelate and hexacyanoferrates was greater at pH 4.0 than at pH 7.0 when heterotrophic fungal cultures were used [11]. Although the metal–cyanide complexes supported fungal growth as nitrogen sources, the rate of growth was only approximately 20% of that observed when HCN was provided as the nitrogen source. Nevertheless, when a mixed culture of F. oxysporum, Sctyalidium thermophilum, and Penicilium miczynski isolated on hexacyanoferrate was grown in the presence of this metal–cyanide complex, >50% loss of cyanide was observed after 28 days of growth. Maximal degradation coincided with maximal growth. Growth of a culture containing F. solani and Trichoderma polysporum on TCN resulted in >90% loss of cyanide in the same time period. When F. solani was grown alone on hexacyanoferrate, there was a 90% loss of cyanide and a 95% loss of iron from the solution in 34 days.

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The conversion of formamide appears to be the rate-limiting step in this process [21]. However, some fungal studies have observed an accumulation of formamide without the accompanying conversion to formate. As the extent to which formate is synthesized by fungi depends upon the culture conditions, there may also be external factors that influence the conversion of formamide. Formation of formate was not observed in washed-cell suspension and cell-free extract systems, but was present in cultures undergoing active growth [21]. Studies showing formation of amidase products were obtained from actively growing cultures [10,68]. Finally, as for bacteria, it appears unlikely that fungi degrade the metal complex itself, but transform the cyanide liberated from the complex as it dissociates in solution. Several factors support this contention. First, the activity of the cyanide hydratase/amidase system is consistent with the fungal mechanism for free cyanide degradation. Since significant structural work has been done on this family of enzymes [39], the ability to accept metal–cyanide complexes substrates appears unlikely. Second, the propensity to degrade metal–cyanide complexes parallels their stability in solution, just as it does for bacteria. Fungal degradation of TCN was more rapid than for hexacyanoferrate. Similarly, measurements of cyanide hydratase and amidase activity were at least tenfold greater for fungi grown on TCN at pH 7.0 than for cultures grown on hexacyanoferrates at pH 4.0. The stability constant for the iron complex is threefold greater the constant for TCN. Furthermore, there was no fungal growth on hexacyanoferrates at pH 7.0, conditions where the dissociation of hexacyanoferrate in solution would be more limited compared to pH 4.0 [69]. Apparent degradation of hexacyanoferrates was also associated with a loss of blue color (Prussian blue) from the nutrient culture [10], which would also be coincident with dissociation of the complex. Nevertheless, little loss of cyanide was observed in sterile systems, indicating that some biological component contributes to the removal of cyanide by fungi in systems containing metal–cyanide complexes.

6.3.3 FUNGAL DEGRADATION OF THIOCYANATE There is limited evidence for the fungal degradation of thiocyanate. Candida tropicalis was identified in a rotating biological contactor used for thiocyanate and phenol treatment, but its specific contribution to degradation was not determined. In another study, Acremonium strictum, isolated from the activated sludge obtained from a wastewater treatment plant for coke-oven-gas condensate, showed >90% degradation of thiocyanate in three days, given an initial concentration of 1.2 g l−1 [55]. Degradation was pH dependent, with an optimum near 6.0, and was influenced by the initial thiocyanate concentration. The maximal rate of degradation was observed at a concentration of 2.1 g l−1 , with the rate decreasing at concentrations >4 g l−1 . Degradation of thiocyanate was inhibited to a limited extent by nitrate and ammonia, but was significantly decreased by elevated concentrations of free cyanide, nitrite, or phenol, again demonstrating the importance of nitrogen sources and cocontaminants in the waste stream on the degradation of cyanogenic compounds. Although the authors of a study with A. strictum concluded otherwise [55], the mechanism of thiocyanate degradation is consistent with the cyanate pathway described in Section 6.1.2.1 for bacteria. The products of thiocyanate degradation by A. strictum were ammonia and sulfate. Although the initial products of the cyanate pathway are cyanate and sulfide, subsequent oxidation of sulfide gives rise to sulfate. Although no cyanate was observed following thiocyanate degradation by A. strictum, there may be several possible reasons for this. Unlike most studies of cyanide degradation, this study was performed upon agar plates, rather than in solution culture. Perhaps these conditions promoted sulfide oxidation by the fungal culture. As there was nearly complete degradation of thiocyanate during experiment, subsequent oxidation to sulfate may have been required to prevent sulfide toxicity. Alternately, there may have been an increased demand for sulfate that may have induced the enzyme responsible for sulfur oxidation in A. strictum. Obviously additional study is required to resolve this question and more thoroughly characterize the mechanism responsible for the fungal degradation of thiocyanate.

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6.4 BACTERIAL ASSIMILATION 6.4.1 BACTERIA CAPABLE OF ASSIMILATING CYANOGENIC COMPOUNDS Bacterial assimilation of cyanide is spread over a wide range of genera, including both Gram-negative and Gram-positive bacteria (Table 6.1). Some of these bacterial species, such as Chromobacterium violaceum and some species of Pseudomonas, are cyanogenic, meaning that cyanide compounds are preferentially synthesized through secondary metabolism (Chapter 3). However, most other bacterial species capable of assimilating cyanide are not cyanogenic. While a primary role of assimilatory pathways is to provide protection from cyanide intoxication, the contribution of cyanide assimilation to primary metabolism is perhaps of greater importance. Several studies have demonstrated that bacteria can utilize cyanide as a source of nitrogen and carbon for growth [7]. Since a wide variety of bacterial, fungal, and plant species synthesize cyanogenic chemicals for release as defensive compounds and deterrents, cyanide in soil and aqueous systems represents a reduced form of carbon and nitrogen that can be more easily and efficiently assimilated that other more oxidized forms of the same elements. It is not surprising then that specific pathways have evolved to provide organisms with the ability to incorporate these compounds directly into primary metabolism. Cyanide assimilation has been studied most extensively in C. violaceum, E. coli, and B. megaterium. Chromobacterium violaceum has been the primary focus [70] principally because this species possess pathways for cyanide both degradation and assimilation [71]. While other species such as Citrobacter freundii and Enterobacter aerogenes may also assimilate cyanide, it has not been clearly established whether this is due to enzymes specific for cyanide assimilation or ancillary activity of other enzymes.

6.4.2 CONDITIONS REQUIRED FOR BACTERIAL ASSIMILATION OF CYANOGENIC COMPOUNDS Unlike the reactions involved in cyanide degradation, evidence of cyanide assimilation has been studied predominantly under in vitro conditions with isolated enzymes. Few studies [71] have attempted to estimate in vivo activity, but the results obtained have been comparable to the in vitro studies. In micro-organisms such as C. violaceum, cyanide assimilation can be detected in vivo under aerobic conditions at the end of the exponential growth phase during periods of active cyanogenesis [71,72]. The activity of enzymes involved in cyanide assimilation, β-cyanoalanine synthase and γ-cyano-α-aminobutyric acid synthase, as well as sulfur transferase enzymes (e.g., rhodanese, 6.6.6.2), tend to increase in cyanogenic bacterial species following periods of cyanogenesis, consistent with their role in preventing cyanide self intoxication [70,72]. β-Cyanoalanine synthase is believed to play the predominant role in the removal of endogenous cyanide produced during highly active periods of cyanogenesis in C. violaceum [71] and in nonmammalian organisms in general [1], primarily because thiocyanate, the product of the reaction catalyzed by rhodanese, is not detected in bacterial cultures [70,71]. The assimilatory reaction mediated by β-cyanoalanine synthase does not require O2 or NAD(P)H [1] and is therefore more energy-efficient than cyanide degradation. Studies of cyanide assimilation by C. violaceum often use growth conditions (addition of glutamate + glycine) that promote cyanogenesis. Growth of C. violaceum on glutamate and glycine induces cyanide production, which consequently results in an increase in the synthesis of enzymes involved in cyanide assimilation [72]. A glutamate + glycine + methionine treatment resulted in a greater removal of cyanide from the media (91%) than cells treated with glutamate alone (62%). Methionine alone has been shown to inhibit cyanide assimilation, as does chloramphenicol, as indicated by a 10% reduction in formation of cyanide derivatives by cells of C. violaceum exposed to chloramphenicol in the presence of glutamate + glycine + methionine.

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Likewise, stationary phase cells of this same species resuspended in solution containing cyanide in conjunction with either glycine or succinate + serine secrete cyanide assimilatory products to the media [73,74]. Evidence also suggests that cysteine plays an important role in cyanide detoxification and perhaps assimilation in bacterial species such as Citrobacter freundii and Enterobacter aerogenes [75]. As discussed in Section 6.4.3.1, cysteine is a cosubstrate for the assimilatory enzymes.

6.4.3 PATHWAYS FOR THE BACTERIAL ASSIMILATION OF CYANOGENIC COMPOUNDS Cyanide assimilation in bacteria is accomplished by substitution reactions carried out by one of three pyridoxal phosphate enzymes: cysteine synthase, β-cyanoalanine synthase, or γ-cyano-αaminobutyric acid synthase. These pathways require a three-carbon skeleton to accept the cyanide ion (CN− ), resulting in the corresponding release of a characteristic ion in exchange. The product of these substitution reactions is the nitrile derivative of an α-amino acid. The production of ninhydrinpositive material (e.g., amino acids) is used as a diagnostic tool to differentiate between cyanide degradation and assimilation [76]. The nitrile produced and the fate of this compound differs between pathways and between bacterial species.

6.4.3.1 Bacterial Assimilation via the β-cyanoalanine Synthase Pathway The β-cyanoalanine synthase pathway is believed to play a primary role in cyanide assimilation in cyanogenic bacterial species like C. violaceum and in acyanogenic species such as B. megaterium and E. coli. The first step in the pathway (Table 6.3) is a pyridoxal phosphate substitution reaction that can reportedly use several potential three-carbon substrates, including cysteine, cystine, serine, O-acetylserine [1,71,73]. This has been observed for enzymes isolated from B. megaterium [77] and Enterobacter strain 10-1 [78,79]. For most bacterial species examined, including E. coli, B. megaterium, and Salmonella typhimurium, cysteine synthase activity is responsible for the formation of β-cyanoalanine. Nevertheless, cyanoalanine synthase (CAS) from C. violaceum has been shown to be the enzyme specifically involved in cyanoalanine synthesis [70]. The subsequent fate of cyanoalanine formed from bacterial CAS and cysteine synthase enzymes has been a matter of conjecture as conflicting reports appear in the literature. In plants, cyanoalanine is converted to the amino acid asparagine (Table 6.3) and subsequently to aspartate with hydrolytic release of ammonia. However, in C. violaceum, amino acids were not detected following the formation of cyanoalanine [71]. Instead, cyanoalanine persisted, which was suggested as evidence that formation of this compound represented a detoxification step, rather than a step toward assimilation. In contrast, most studies have reported that assimilation by C. violaceum, E. coli, and B. megaterium, parallels plant assimilation with asparagine or aspartate detected subsequent to cyanoalanine formation [1,70,71,73,80–83]. The enzyme responsible for this reaction seems to differ by bacterial species. β-Cyanoalanine hydratases are responsible for this reaction in Pseudomonas sp. [84], and perhaps also in B. megaterium and C. violaceum. Chromobacterium violaceum also has an alternate enzymatic system for cyanide assimilation (Section 6.4.3.2). The enzyme from Pseudomonas sp. strain 13, which catalyzes the formation of asparagine and aspartate is a large 100 kDa protein specific for cyanoalanine, yielding the products asparagine and aspartate in a 2.2:1 ratio. A comparable cyanoalanine hydratase has not been characterized from any other bacterial source. In contrast, for Endobacter strain 10-1 and E. coli, this reaction reportedly occurs through ancillary activity of asparaginase, converting cyanoalanine to asparagine or aspartate [76,82].

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TABLE 6.3 The General Categories of Chemical Reactions Responsible for the Assimilation of Free Cyanide by Living Organisms Substitution reactions ␤-Cyanoalanine synthase HS9CH2CH(NH2)COOH + HCN → H2S + NC9CH2CH(NH2)COOH H3C9COO9CH2CH(NH2)COOH + HCN → CH3COO– + NC9CH2CH(NH2)COOH

H2NCO–CH2CH(NH2)COOH

␥-Cyano-␣-aminobutyric acid synthase – 9 S (CH2)2CH(NH2)COOH + CN– → HS9(CH2)2CH(NH2)COOH + NCS9(CH2)2CH(NH2) → NC9(CH2)2CH(NH2)COOH + SCN– Sulfur transferases S2O32– + CN– → SO32– + NCS–

Amino acid synthesis reactions α-Aminobutyric acid synthesis CH3 CH2 CHO + NH3 + HCN → CH3 CH2 CH(NH2 )–CN → CH3 CH2 CH(NH2 )COOH Alanine synthesis + CH3 CHO + NH+ 4 + HCN → CH3 CH(NH2 )–CN + 2 H2 O → CH3 CH(NH2 )COOH + NH4 Glutamate synthesis HOOC(CH2 )2 CHO + NH+ 4 + HCN → CH3 CH2 CH(NH2 )–CN + 2 H2 O → HOOC(CH2 )2 CH(NH2 )COOH + NH3

For the β-cyanoalanine pathway, there are two possible substrates, which can accept free cyanide. The synthesis of γ-cyanoα-aminobutyric acid from free cyanide uses the ionized form (− S–) of homocysteine. This reaction also forms thiocyanate (SCN− ), as do reactions mediated by sulfur transferase enzymes. As α-aminobutyric acid is a possible amino acid precursor, this reaction is included here.

6.4.3.2 Bacterial Assimilation via γ-Cyano-α-aminobutyric Acid Synthase An alternate pathway for cyanide assimilation is apparently unique to C. violaceum [85]. This pathway, which also requires pyridoxal phosphate, operates in parallel with the β-cyanoalanine synthase and is induced by similar growth conditions, such as growth on media containing glutamate + glycine [74]. As mentioned earlier, this pathway represents only a minor pathway for assimilation, with the γ-cyano-α-aminobutyric acid possibly serving as a precursor for the synthesis of amino acids such as glutamate. The synthesis of γ-cyano-α-aminobutyric acid from cyanide occurs via a two-step reaction (Table 6.3), the first of which is nonenzymatic and the second catalyzed by γ-cyano-α-aminobutyric acid synthase. The second reaction is fundamentally an enzymatic replacement reaction for thiocyanate (SCN). The available data demonstrates that γ-cyano-α-aminobutyric acid synthase is highly specific in its enzymatic activity. This enzyme shows no comparable activity as a homoserine sulfydrylase or OAS homoserine sulfhydrylase synthesizes cystathionine at a low rate (8%) in comparison to the formation of γ-cyano-α-aminobutyric acid. The importance of this amino acid derivative and this pathway have not been established.

6.5 FUNGAL ASSIMILATION 6.5.1 FUNGI CAPABLE OF DEGRADING CYANOGENIC COMPOUNDS Fungal assimilation of cyanide has a more limited distribution than is observed in bacteria and plants. Assimilation has been observed in as few as five genera, representing only eight species in total. Three species capable of assimilating cyanide appear in the genus Pholiota, and include P. adipose,

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P. aurivella, and P. praecox. Another species, identified as a psychrophilic basidiomycete and the pathogen responsible for winter crown rot in forage crops, is also capable of cyanide assimilation [86]. This species may be a snow mold fungus, perhaps Fusarium nivale. Previous studies have shown F. nivale to be cyanogenic [87], as well as a species capable of cyanide assimilation. Two other genera that reportedly assimilate cyanide, Rhisopus and Marasmius, are also clearly cyanogenic. Rhizopus oryzae, a fungus associated with cassava spoilage, reportedly detoxifies cyanide with extracellular rhodanese [88], but the contribution of this process to cyanide assimilation is unclear and may represent detoxification rather than assimilation.

6.5.2 CONDITIONS REQUIRED FOR THE FUNGAL ASSIMILATION OF CYANOGENIC COMPOUNDS Fungal assimilation of cyanide has been observed in cultures grown in standard media formulations at both room temperature and decreased temperature (13◦ C). Synthetic media have been used in most cases [86,89], with potato dextrose used in one case where synthetic media failed to support growth of specific fungal strain [90]. Similar to bacteria, assimilation is observed when specific substrates are present in the media when cyanide is present. For fungi, assimilation has been observed when ammonia and acetaldehyde are supplied simultaneously with cyanide [91], or in the presence of ammonia and succinic semialdehyde [92]. The additional substrates are required for the specific cyanide assimilatory pathways present in fungi, both of which lead to the synthesis of specific amino acids.

6.5.3 PATHWAYS FOR THE FUNGAL ASSIMILATION OF CYANOGENIC COMPOUNDS As with plants and bacteria, the fungal pathways for cyanide assimilation are associated with amino acid metabolism. One of these pathways, present in F. nivale, is comparable to the β-cyanoalanine pathway in that the product is asparagine. Little is known about this pathway other than the product formed, so this pathway will not be discussed in detail here. Previous studies have shown that β-cyanoalanine synthase is not the enzyme responsible for cyanide detoxification in fungal species such as Stemphylium loti [93]. A second pathway, similar to that found in bacteria, has been described in Rhizoctonia solani and involves an ammonia-dependent conversion of cyanide to α-aminobutyric acid [89]. This product of this pathway reportedly differs from the corresponding bacterial pathway in the carbon atom onto which the cyanide moiety is substituted and the substrates required. Most fungi capable of assimilating cyanide utilize a second pathway leading to the formation of alanine, including the aforementioned psychrophilic basidomycete. This species can also form glutamic acid via a similar pathway. Production of both amino acids in response to cyanide exposure has been reported in one study [86]. 6.5.3.1 Fungal Assimilation via α-Aminobutyric Acid Synthesis Rhizoctonia solani is a pathogenic fungus that causes root rot in crop plants. In the presence of cyanide and ammonia, this species forms the intermediate α-aminobutyronitrile from proprionaldehyde (Table 6.3). The proprionaldehyde is believed to arise from the nonoxidative decarboxylation of α-ketobutyrate [94]. Studies of cyanide assimilation in liquid cultures of R. solani at room temperature demonstrated rapid formation of α-aminobutyronitrile within 30 min, followed by the appearance of α-aminobutyric acid with the next 30 min. This observation was interpreted as indicative of a precursor–product relationship [89]. This second step requires water, resulting in an enzymatic hydrolysis to product within the regeneration of one molecule of ammonia. Presumably, the resulting α-aminobutyric acid can be used to synthesize amino acids. The α-aminobutyric acid formed differs from that produced by the corresponding bacterial pathway in that the cyanide is added to the aldehyde carbon atom along with the ammonia

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molecule. The subsequent hydrolysis reaction converts the cyanide group to a carboxyl group, resulting in the complete assimilation of cyanide in these two steps. In the bacterial pathway, the cyanide is incorporated into α-aminobutyric acid via substitution onto a sulfhydryl group at the terminal carbon on the opposite end of the molecule. The cyanide molecule persists on the product γ-cyano-α-aminobutyric acid. 6.5.3.2 Alanine and Glutamate Synthesis The pathways for alanine and glutamate synthesis in a previously unidentified physchrophilic basidiomycete are similar to the α-aminobutyric acid pathway above in that ammonia is required as a precursor and the second hydrolytic step regenerates ammonia. The pathways in this basidiomycete also lead directly to amino acids, rather than an amino acid precursor. Formation of alanine utilizes acetaldehyde as the precursor (Table 6.3). The ubiquity of acetaldehyde as an intermediate in cellular metabolism perhaps explains why the majority of fungal species capable of cyanide assimilation produce alanine following exposure to cyanide [91]. Little kinetic information has been reported for this proposed reaction. The synthesis of alanine by this pathway occurs at a somewhat slower rate than α-aminobutyric acid. Fungal cultures exposed to cyanide showed a rapid increase in α-aminopropionitrile during the 12 h following exposure, followed by a subsequent decrease over the next 12 h to initial levels. In contrast, alanine levels increase steadily over the same 24 h period. While there is a linear relationship between formation of αaminopropionitrile formation and enzyme (as total protein) concentration, α-aminopropionitrile also formed nonenzymatically when the protein extract was omitted from the reaction [91]. The enzyme responsible for the formation of alanine has not been described, but the reaction is reportedly similar to a plant nitrilase reaction. The initial studies of cyanide assimilation in this fungal species demonstrated initial formation of both alanine and lower levels of glutamate [86]. This observation led to the hypothesis that the glutamate was formed from alanine [91]. Subsequent study revealed, however, that succinic semialdehyde can be converted to 4-amino-4-cyanobutyric acid in the presence of ammonia and cyanide (Table 6.3). As for the formation of alanine, crude isolated protein extracts of the fungal mycelia showed evidence of nitrilase activity [92]. Nitrilase activity showed a linear increase with time and enzymes concentration at room temperature, with a pH optimum of 8.0. Similar protein extracts also possessed other enzymatic activity, including glutamic acid decarboxylase, succinic semialdehyde dehydrogenase, and γ-aminobutyric acid transaminase. Based upon these various activities, a pathway was proposed that relates formation of 4-amino-4-cyanobutyric acid to basic elements of fungal carbon (succinic semialdehyde) and nitrogen (glutamate) metabolism (Figure 6.1).

6.6 BOTANICAL TRANSFORMATION PROCESSES Cyanide is involved in a wide variety of plant pathways (Chapter 3). Free cyanide (CN− ) is produced concomitantly with the gaseous plant hormone ethylene and is the active constituent of the cyanogenic glycosides used to deter herbivory and store nitrogen, and may regulate aspects of nitrogen metabolism. Another cyanide compound, cyanate (CNO− ), is produced from the decomposition of urea and other compounds. Not surprisingly, there are corresponding pathways that control endogenous levels of these and other cyanide compounds in plant tissues, preventing self-intoxication. Despite periods of active cyanogenesis, concentrations of free cyanide are typically maintained at low levels in plants [95]. There are two primary pathways that contribute to the assimilation of cyanide compounds. Plants appear not to have pathways for cyanide degradation. Free cyanide (HCN, CN− ) is removed by either the cyanoalanine pathway (Section 6.4.3.1) or the sulfur transferase pathway (Section 6.6.2.2). The presence of a plant sulfur transferase pathway in plants similar to that of rhodanese in animals was long disputed but was recently confirmed when genes encoding proteins

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Succinic semialdehyde NH4+ Ketoglutaric acid HCN

gAminobutyric acid 4-Amino 4-cyanobutyric acid

2H2O

CO2

Glutamic acid

NH4+

FIGURE 6.1 A biochemical model illustrating the assimilation of free cyanide into 4-amino-4-cyanobutyric acid by a psychrophilic basidiomycete (fungi). The 4-amino-4-cyanobutyric acid is converted to glutamate by the action of a potentially unique nitrilase. (Source: Strobel, G.A., J. Biol. Chem., 242, 3265, 1967. With permission.)

with sulfur transferase activity were identified. The specific role of the sulfur transferase pathway and cyanase in plant metabolism is still largely a matter of conjecture. The relative importance of these pathways with respect to one another and to cyanide assimilation/detoxification in plants is also unknown. The interaction between these pathways has only been examined in insects, where the cyanoalanine pathway has been shown to play the primary role [96]. Other enzymes, such as cyanase and cyanamide hydratase, contribute to the metabolism of other natural and anthropogenic cyanide compounds. The role of these enzymes in plant metabolism is not clear, although roles for cyanase in primary carbon metabolism have been proposed [29].

6.6.1 CYANIDE-RESISTANT RESPIRATION Cyanide disrupts numerous metabolic processes, principally by binding to the metal cofactors in enzymes, although reactions with functional groups such as carbonyls or disulfide bonds also occur. The most detrimental inhibition is blockage of the mitochondrial electron transport chain, through complexation of cyanide with the iron (Fe3+ ) in the terminal cytochrome, cytochrome c oxidase. In animals, exposure to cyanide can quickly decrease respiration rates to <1% of the initial level. Inhibition of this cytochrome uncouples ATP synthesis from electron transport, depleting the cellular ATP pool, and causing a shift to anaerobic metabolism, specifically through the pentose phosphate pathway. Cyanide toxicity is related to dependence of the organism on aerobic metabolism. Plants and algae are protected to a degree from cyanide poisoning because their mitochondrial electron transport chain includes an alternative oxidase (Figure 6.2). While not directly associated with cyanide transformation, this protein does contribute to a plant’s ability to resist cyanide toxicity and thus deserves mention here. The plant mitochondrial alternative oxidase transfers electrons from the ubiquinone pool to oxygen, producing water and heat [97], allowing electron transport rates of 10 to 100% of the uninhibited control rate. While the diversion of electrons to the alternative oxidase does not produce ATP, the first proton-transporting cytochrome is not bypassed. A proton motive force for ATP synthesis is still generated, albeit at a reduced capacity [98]. While the alternative oxidase is not involved in

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NAD(P)H

Intermembrane space

NAD(P)

cyt cc cyt

UQ

II

II

III III

IV

AOX 1 2 O2

NAD(P)H

NAD(P)

Succinate

Fumarate

H2O

Matrix

FIGURE 6.2 A schematic representation of the plant mitochondrial electron transport chain, illustrating the diversion of electrons through the alternative oxidase (AOX) in the absence (solid line) or presence (dotted line) of cyanide. Electrons from NAD(P)H are fed into the ubiquinone (UQ) pool via Complex I or via the two inner or two outer NAD(P)H dehydrogenases. In the absence of cyanide, electrons flow from the UQ pool through Complex II, Complex III, cytochrome c, and Complex IV, with the electrons accepted by oxygen to form water. In the presence of cyanide, electron flow through Complex IV is prevented. Under these conditions, electron flow is diverted after Complex II to the AOX, and then subsequently to oxygen.

the detoxification of cyanide per se, this pathway does serve to mitigate adverse effects of cyanide on the mitochondrial electron transport chain, protecting this important biochemical system until the potential for detoxification is realized. Cyanide-resistant respiration may serve a general role by preventing over-reduction of the ubiquinone pool and by reducing free radical production when respiratory enzymes are not keeping pace with electron transport, such as occurs in response to tissue damage or physiological stress, including temperature, drought, [98] and exposure to cyanide.

6.6.2 PLANT PATHWAYS FOR THE ASSIMILATION OF CYANOGENIC COMPOUNDS 6.6.2.1 Plant Assimilation via the Cyanoalanine Pathway The cyanoalanine pathway has been postulated to mediate the removal of cyanide produced by ethylene synthesis or hydrolysis of cyanogenic glycosides [95,99,100]. As in bacteria, the first step in the pathway utilizes the amino acid cysteine, replacing the thiol group with cyanide to form cyanoalanine (Table 6.3). In plants, cyanoalanine synthase activity, the first enzyme in the pathway, is modulated during pathogenic attack, stress [101,102], and by ethylene [103–109]. Activity is also present in potato tubers [109] and both imbibed and dry seeds of cocklebur, rice, and soybean [106,107,110]. CAS is believed to have a fundamental role in amino acid metabolism [107]. As cyanide reversibly inhibits nitrate reductase, CAS activity is likely involved in the modulation of nitrogen assimilation. The activity of CAS has also been related directly to an increase in the amino acid pool of seeds, specifically to asparagine and aspartate [107]. The second step in this pathway converts cyanoalanine to asparagine (Table 6.3). Cyanoalanine can also be present in plant tissues as a γ-glutamyl dipeptide. β-Cyanoalanine hydrolase catalyzes

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the formation of asparagine and shows a high specificity for β-cyanoalanine, displaying no activity with respect to other amides and nitriles, including asparagine, arginine, glutamine, nicotinamide, ricinine, or the auxin precursors indole acetamide and indole acetonitrile [80]. Nitrile aminohydrolase from Arabidopsis thaliana can also catalyze this step [111] and is considered a member of the NIT group of plant nitrilases. Nitrilase 4 differs from nitrile aminohydrolase in that it is encoded by a single gene located on a different chromosome and, like β-cyanoalanine hydrolase, does not accept indole acetonitrile as a substrate. Both asparagine and aspartate can be products of this pathway. The mechanism by which nitrilase 4 discriminates between nitrilase and nitrile hydratase activity is unclear. The activity of the cyanoalanine pathway has been directly related to cyanide tolerance in plants. Recent data has demonstrated that herbicides such as quinolinecarboxylic acid (Quinclorac) and 2,4-dichlorophenoxyacetic acid (2,4-D) stimulate ethylene biosynthesis [110,112–114] because they are structural analogs of plant auxins. Auxins are the principal growth-regulating compound in plants. Quinclorac is an auxinic herbicide used against barnyard grass (Echinochloa crus-galli) and other grasses. The herbicidal mode of action is related to HCN production during ethylene synthesis [110]. Increases in ethylene, HCN, and CAS activity have been detected in barnyard grass following quinclorac application. Quinclorac tolerance in rice is associated with a higher activity of CAS [114]. The herbicide 2,4-dichlorophenoxyacetic acid also increased CAS activity in soybeans [113]. The response of the CAS pathway to the manipulation of natural cyanide production suggests that this pathway may also detoxify anthropogenic cyanide contamination in the environment (Chapter 24). Several plant species, including basket willow (Salix viminalis), white willow (Salix alba), Balsam poplar (Populus trichocarp), elderberry (Sambucus sp.) [115,116], and diamond willow (Salix eriocephala) [117] are capable of transporting and metabolizing exogenous free cyanide. For the study with diamond willow, the stable isotope 15 N was used to track the movement of the cyanogenic nitrogen atom. Analysis of the amino acid fraction extracted from roots and leaves of cyanide-treated willow plants demonstrated that the 15 N was incorporated into all amino acids, but with the greatest enrichment observed in asparagine, consistent with activity of the cyanoalanine pathway [118].

6.6.2.2 Plant Assimilation via the Sulfur Transferase Pathway A second pathway in plants for cyanide detoxification involves sulfur transferase (ST) enzymes (Table 6.1), similar to the detoxification of cyanide in animals catalyzed by rhodanese. The hallmark of ST activity (rhodanese, thiosulfate:cyanide transferase, or 3-mercaptopyruvate sulfur transferase) is the transfer of sulfur from thiosulfate to cyanide, forming thiocyanate (Table 6.3). Although rhodanese has been demonstrated in the green alga Chlamydomonas reinhardtii [119], the corresponding gene in plants was only recently obtained when two cDNAs from A. thaliana encoding enzymes with rhodanese-like activity were characterized [120]. Additional studies have identified A. thaliana genes with sulfur transferase activity, catalyzing the transfer of sulfur from thiosulfate to cyanide [121,122]. Two other genes were also isolated, purportedly encoding mercaptopyruvate sulfur transferase enzymes with homology similar to the A. thaliana genes [123]. Although these genes have been identified, the role of this pathway in plants has not been established. A possible role in senescence has been suggested because ST activity and steady state RNA levels increase with plant age for A. thaliana [122].

6.6.2.3 Plant Assimilation via Cyanase Cyanate (CNO− ) can also be involved in plant cyanide metabolism. Cyanate can form spontaneously in cyanide-containing mining wastes, photochemically from metal–cyanide complexes, or naturally

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from the dissociation of urea or the decomposition of carbamoylphosphate [124]. Carbamoylphosphate is a precursor of nucleotides formed from the reaction of glutamine with bicarbonate, catalyzed by carbamoyl phosphate synthetase. Carbamoylphosphate production can also result from the inhibitory activity of some plant pathogens. For example, Pseudomonas syringae produces a tripeptide called phaseolotoxin, which inhibits arginine synthesis and can lead to increased production of carbamoylphosphate [125]. Carbamoylphosphate is predicted to have a short half-life in aerial plant tissues (<10 min) at ambient temperature and will readily form cyanate. Cyanate, as a nucleophile, attacks amino and thiol groups, leading to inhibitory effects as concentrations increase. Cyanate production increases in plant tissue with temperature, resulting in what could be considered heat-induced cyanate stress [124]. Cyanate can also be absorbed directly into plant tissues when present in foliar herbicides. In addition to possible nucleophilic reactions, cyanate can react with dehydroascorbate to form carbamoyldehydroascorbate [126,127], which has been detected at concentrations as high as 3.8 µmol kg−1 FW in potato tissues. This reaction between cyanate and dehydroascorbate upsets the balance between ascorbate and dehydroascorbate, leading to a depletion of ascorbate and the development of oxidative stress. Cyanate is removed enzymatically through decomposition catalyzed by cyanase, a bicarbonatedependent conversion to ammonia and carbon dioxide that is similar to that observed in bacteria (Table 6.2). While the principal role of this reaction is the detoxification of cyanate, the importance of cyanase activity may be more closely linked to primary metabolism [124]. Degradation of cyanate releases ammonia that can be used in nitrogen metabolism, similar to the cyanoalanine pathway. Cyanase may also provide a means of concentrating and delivering carbon dioxide to photosynthetic cells, acting in concert with carbonic anhydrase to regulate cellular carbondioxide/bicarbonate balance. Likewise, cyanase may be important to the ascorbate/dehydroascorbate balance described earlier and to cellular pH, as the reaction catalyzed by cyanase utilizes protons. These roles in primary metabolism are speculative and require additional scientific study. 6.6.2.4 Plant Assimilation of Cyanamide Herbicides Cyanamide was used extensively in agriculture during the 20th century because this compound acts as both a preemergence fertilizer and fungicide and as a postemergence herbicide and defoliant. The introduction of other nitrogen fertilizers resulted in a decline in cyanamide use, as did evidence of carcinogenicity. However because of the reactivity of the nitrile group on this molecule, use of cyanamide increased in industry. Cyanamide is used currently for the production of pig iron and in the chemical synthesis of plastics, resins, urea, guanidine, and other compounds. Plants treated directly with cyanamide at rates comparable to those employed in the field suffer rapid defoliation and toxicity. Plants treated with cyanamide at lower rates (e.g., 2.4 mM) are reportedly capable of metabolizing the herbicide, particularly if treated simultaneously with ornithine (Equation [6.1]). HOOCCH(NH2 )(CH2 )2 CH2 NH2 +H2 NCN → HOOCCH(NH2 )(CH2 )3 NHCNH(NH2 ) ornithine cyanamide arginine (6.1) The amino acid arginine is synthesized from ornithine. Plants treated with cyanamide reportedly showed an increase in arginine content [128]. Roots of rape (Brassica napus) from plants treated with cyanamide in conjunction with ornithine showed a 30-fold increase in arginine levels as compared to roots from plants treated with nitrate and a fourfold increase as compared to roots from plants treated with nitrate and ornithine. A similar pattern was observed for leaves, with cyanamide + ornithine treated plants displaying a 12-fold higher arginine content than plants treated with either nitrate or nitrate+ornithine. Tissue lysine and histidine levels showed a two- to eightfold increase as compared to the same controls, purportedly indicating plant metabolism of cyanamide to lysine and histidine,

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followed by the formation of arginine. This reportedly involved the direct formation of arginine, rather than via formation of citrulline and was based upon data demonstrating a decrease in arginase activity in plants treated with cyanamide. The natural but perhaps limited ability of plants to metabolize cyanamide can be enhanced through biotechnology. Transgenic plants capable of tolerating and metabolizing cyanamide were engineered by expressing a fungal cyanamide hydratase gene (cah) in tobacco [129]. This enzyme hydrates cyanamide to urea. Tobacco plants expressing the cah gene were capable of growth in media containing 2.4 to 12 mM cyanamide, concentrations that were lethal to control plants. An accumulation of urea in transgenic plants was also reported, providing additional evidence for activity of the cah gene in cyanamide metabolism. 6.6.2.5 Plant Transport and Transformation of Metal–Cyanide Complexes Study of the plant-mediated degradation of metal–cyanide complexes is a relatively new area of research, primarily because large complexes such as ferricyanide and ferrocyanide were believed to be membrane impermeable [130]. Several studies have examined the degradation of iron cyanide complexes by bacteria and fungi and generally suggest that biodegradation, if it occurs at all, is limited [24] and must occur under aerobic conditions [131]. A small number of bacterial [45,46] and fungal [10,11] strains have been isolated that reportedly use cyanogenic nitrogen from metal– cyanide complexes as the sole source for growth. In several cases, these organisms were isolated from cyanide-contaminated media. Given the evolutionary relationship between cyanide metabolic pathways in micro-organisms, fungi, and plants, these studies imply that the genetic potential for plant biodegradation and assimilation of metal–cyanide complexes may also exist. The only published report of ferrocyanide degradation by a plant involved a study which utilized a willow (Salix eriocephala var Michaux) clone found growing on the fringe of iron cyanide contamination at a former manufactured gas plant site [117]. The transport and metabolism of two stable isotope-labeled (15 N) cyanide compounds, potassium cyanide and potassium ferrocyanide were examined. This study used a modified nutrient solution developed using a chemical equilibrium model [132] and the most recent thermodynamic data for ferrocyanide dissolution [69,133] to maintain the added cyanide in soluble form and preserve the initial speciation. A novel tissue extraction procedure was further used to extract cyanide compounds and relate the in vivo tissue cyanide content and speciation tissue to tissue 15 N content [134]. Analysis of plant tissues following a 20 day exposure to potassium ferrocyanide, added at a rate equivalent to 2 ppm total cyanide, revealed that only trace amounts of cyanide were detected in the aerial tissues of plants. Highly significant increases in 15 N enrichment were observed, however, suggesting transport and subsequent metabolism of ferrocyanide. Mass balance calculations for tissue 15 N and solution cyanide confirmed 100% recovery of the added ferrocyanide. While there was some dissociation of ferrocyanide in the nutrient solution, as determined from analyses of solution samples, the total mass of free cyanide released was insufficient to account for the increase in 15 N observed. A biodegradation study utilizing micro-organisms from the hydroponic solution indicated that the bacteria present could not degrade the added ferrocyanide. The results of this study are highly suggestive of ferrocyanide transport and metabolism in plants, augmenting earlier reports of plant transport of metal–cyanide complexes [116,135]. Additional study will be required, however, to determine the mechanism of plant-mediated biodegradation of ferrocyanide and the metabolites produced by this biodegradation. Direct biodegradation of the complex may not be feasible, at least by the enzymes described earlier. If complexed cyanides cannot be degraded directly, then metal–cyanide complexes may have to dissociate first to free cyanide, followed by activity of one of the metabolic pathways described earlier. Phytophotolysis (light-mediated dissociation within plant tissues) may play a role, as it does for some organic contaminants, but this presumes transport of the complexed cyanide to leaves.

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There is circumstantial evidence that plants may form complexed cyanides in foliar tissues when exposed to free cyanide. Studies with various species of willow have revealed the presence of a chemical species of cyanide that behaves as a strong cyanide complex [116,117]. When the extraction approach described earlier [134] was used on willow tissues from plants exposed to potassium cyanide [117], the results revealed that cyanide in the leaf tissues could be detected only if the distillation protocol for total cyanides was utilized [136], suggestive of the formation of strong cationic complexes within plants. Another possibility is that both free and complexed cyanide are transported to aerial tissues [116]. The possible processes involved in the formation of these strong cyanide complexes have not been investigated. 6.6.2.6 Volatilization of Cyanide by Plants The plant transformations of cyanide discussed thus far have considered biochemical transformations. Another potential pathway for cyanide transformation by plants is the conversion of cyanide from a dissolved chemical species within cells to a volatile form. Plants regularly release gaseous cyanide during ethylene synthesis and during the hydrolysis of cyanogenic glycosides. Of additional interest, particularly from a remediation and regulatory perspectives, is the possible volatilization of exogenous cyanide by plants. The available data on this subject is somewhat contradictory, so additional research is clearly warranted. Studies with various species of willow have provided evidence of cyanide volatilization. Draeger tubes placed in proximity to willow (S. eriocephala var. Michaux) plants exposed to a solution containing a total of 6 mg l−1 iron cyanide in a 95:5 ferrocyanide:ferricyanide ratio [137] showed evidence of exposure to volatile cyanide gas. Experiments with other species of willow plants exposed to free cyanide also reported cyanide volatilization, with the extent dependent upon light and transpiration [115]. However, cyanide volatilization is believed to represent only a small fraction of the total cyanide mass that initially moved into the plant tissue. In contrast, no volatilization of cyanide was observed for willow (S. eriocephala) [117] or by pea (Pisum sativum) [138]. Two separate techniques, picrate papers [139] and plant enclosures with controlled airflow, were used, neither of which provide evidence of volatilization. A mass balance calculated for willows exposed to free cyanide achieved only a 60% recovery [117]. Since the experimental system used a closed hydroponic system, one possibility is that the difference is due to the loss of volatile cyanide. However, a model [134] developed using the data from this experiment predicted that volatilization from these plants was negligible. Additional study is required to more comprehensively evaluate the possible volatilization of cyanide by plants.

6.7 ENVIRONMENTAL IMPACTS OF CYANIDE ASSIMILATION The ecological relevance of the cyanide degradation and assimilatory pathways in bacteria, fungi, and plants has been a matter of conjecture. Certainly these pathways provide a means of cyanide detoxification, preventing self-intoxication in cyanogenic organisms, or as a means of countering cyanogenesis as a biochemical defense. Assimilation has the benefit of providing a means of using cyanide to form primary metabolites such as amino acids or precursors. As described in Chapter 3, cyanide is a common compound in the environment, produced by a wide variety of cyanogenic organisms, including bacteria, fungi, plants, and some animals. With this wide array of cyanide-producing organisms, encountering cyanide in terrestrial and aqueous environments would not perhaps be wholly unexpected, although the concentration would likely vary considerably, particularly in anthropogenic waste streams. While this anthropogenic cyanide represents a potential metabolic inhibitor that must be detoxified if present at elevated concentrations, this molecule may represent a nutritionally-available form of carbon and nitrogen utilized by multiple organisms.

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The natural presence of cyanide in the environment has been postulated to be a component of cyanide “microcycles” [90,140]. These microcycles would consist of cyanogenic organisms and organisms that assimilate cyanide as a source of carbon and nitrogen for growth. The role of cyanide in such ecosystems would depend upon the species composition and distribution. The nature of the interactions between these organisms might also vary. Cyanogenic organisms and cyanide-assimilating organisms could potentially be involved in symbiotic, commensal, or pathogenic relationships. For example, sorghum is a cyanogenic plant that produces cyanogenic glycosides as a defensive compound. Sorghum can become infected with the pathogen Gloeocercospora sorghi, a fungi capable of detoxifying cyanide. One possible scenario is that the fungal pathogen may elicit cyanide release by sorghum from cyanogenic glycosides [140]. Since the fungus can metabolize cyanide, this pathogen would feed upon the cyanide released, assimilating this molecule as a source of carbon and nitrogen to continue growth and advance the infection. This scenario is just one example of a community in which cyanide plays a specific ecological role. Unfortunately, there has been little work in this area beyond the study of synthesis, assimilation, and degradation pathways in these organisms. The relative importance of cyanide microcycles in the environment remains unclear. In contrast, the contribution of these natural pathways to the processing and remediation of cyanogenic compounds present in various waste streams is well documented. These contributions are discussed in greater detail in Chapters 23, 24, 26, and 27.

6.8 SUMMARY AND CONCLUSIONS • Bacteria, fungi, and plants have a variety of pathways capable of transforming cyanogenic compounds, including free cyanide, nitriles, thiocyanate, and metal–cyanide complexes. These pathways either involve the degradation of cyanogenic compounds or their direct assimilation into primary metabolism. The degradation of metal–cyanide compounds most likely involves transformation of free cyanide liberated as the complexes dissociate. • Degradation pathways, which can be oxidative, reductive, or hydrolytic, convert cyanide to simple molecules such as formamide, formic acid, ammonia, and carbon dioxide. Degradation pathways are present in bacteria and fungi, but not in plants. • The degradation of cyanogenic compounds is influenced by a number of factors, including pH, the presence of carbon and nitrogen sources in the media, the availability of oxygen, the presence of cocontaminants in the waste stream, the bioavailability of the cyanogenic compound, and the initial concentration of the cyanogenic compound. • Assimilatory pathways, such as the cyanoalanine pathway, incorporate cyanide into primary metabolites, generally amino acids or their derivatives. Alternately, cyanide can be assimilated into thiocyanate. All plants, as well as a limited number of bacteria and fungi, carry out cyanide assimilation. • The biological pathways for cyanogenesis (cyanide synthesis) and cyanide transformation form microcycles in the environment that allow for the cycling of the carbon and nitrogen atoms from the cyanide molecule. The ecological importance of cyanide microcycles has not been studied in detail. • The range of pathways for cyanide transformation have allowed for the development of a variety of cyanide biotreatment and bioremediation systems, including phytoremediation.

REFERENCES 1. Solomonson, L.P., Cyanide as a metabolic inhibitor, in Cyanide in Biology, Vennesland, B., Conn, E., Knowles, C., Westley, J., and Wissing, F., Eds., Academic Press, London, 1981, p. 11. 2. Pettet, A.E.J. and Mills, E.V., Biological treatment of cyanides with and without sewage, J. Appl. Chem., 4, 434, 1954.

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7 Analysis of Cyanide in Water Rajat S. Ghosh, David A. Dzombak, Sharon M. Drop, and Anping Zheng CONTENTS 7.1 7.2

Interferences and Pretreatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Cyanide Measurement Techniques. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.1 APHA/AWWA/WEF (Standard Methods), ASTM, USEPA, and USGS Methods. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.2 Analytical Finish Techniques. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3 Weak Metal–Cyanide Complexes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.1 Weak Acid Dissociable (WAD) Cyanide Method . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.2 Available Cyanide by Ligand Displacement Method . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.3 Cyanide Amenable to Chlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4 Metal–Cyanide Complex Analysis by Liquid Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.1 Ion Exchange Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.2 Reversed-Phase Ion-Pair Partition Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5 Free Cyanide Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.1 Gas Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.2 Direct Colorimetric Development . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.3 Gas–Liquid Diffusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.4 Ion-Selective Electrode . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6 Thiocyanate and Cyanate Measurement Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.1 Thiocyanate Measurement Procedure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.2 Cyanate Measurement Procedure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.7 Cyanogen Halide Analysis. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.8 Organocyanide Measurement Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.9 Comparative Method Performance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.10 Quality Control Criteria for Measurement of Cyanide in Water . . . . . . . . . . . . . . . . . . . . . . . . . . 7.11 Measuring Cyanides — A Regulatory Dilemma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.12 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

127 128 129 131 132 132 133 133 134 134 136 137 137 137 138 141 141 141 142 143 143 144 148 149 149 150

Dissolved phase cyanide can exist in various inorganic and organic forms. Inorganic forms of cyanide are usually classified into three forms, namely, (i) free cyanide, (ii) weak metal–cyanide complexes; and (iii) strong metal–cyanide complexes [1]. The simplest and the most reactive class of cyanide species is free cyanide, which includes both the cyanide anion, CN− , and its protonated form HCN(aq). The next most reactive species are the soluble complexes of the cyanide ion with certain transition metals like copper, zinc, cadmium, nickel, and silver. This group of transition metal– cyanide complexes is commonly known as weak metal–cyanide complexes or weak acid dissociable cyanides because of their ability to release the cyanide ion in aqueous solutions under weak acid 123

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TABLE 7.1 List of Common Dissolved Cyanide Species Associated with Contaminated Waters Dominant dissolved species

Classification of dissolved species

CN− (pH > 9.2), HCN (pH < 9.2) 2− − Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , 2− − Ni(CN)4 , Ag(CN)2 4− 3− Fe(CN)3− 6 , Fe(CN)6 , Co(CN)6

HO

HO

CH2 HO

Free cyanide Weak acid dissociable

Strong acid dissociable

O

HO HO

CH2OH O O

OCH2

CH3

O CH

HO

HO

CN C

C N

(Acetonitrile)

OCH

(Amygdalin)

HO

CH3

CH3

OH OH (Dhurrin)

FIGURE 7.1 Molecular structures of common organocyanides.

conditions. Finally, there are strong cyanide complexes with other transition metals, for example, gold, platinum, cobalt, and iron. Because of strength of the metal–cyanide bonds in these complexes, they often require strong acidic pH conditions (pH < 2) and heat to dissociate and release the cyanide ion. As a result, these complexes are often referred to collectively as strong metal–cyanide complexes or strong acid dissociable cyanides. Table 7.1 lists some common inorganic cyanide species encountered in contaminated waters. A more complete list is provided in Chapter 5. Other inorganic forms of cyanide that are sometimes of interest in water are thiocyanate, SCN− , cyanate, CNO− , and cyanogen halides such as cyanogen chloride, CNCl. None of these species is captured in the conventional total cyanide test. Thiocyanate occurs in various kinds of industrial wastewaters [2], but is not as common a cyanide contaminant as the free cyanide species and metal–cyanide complexes. As discussed in Chapter 5, cyanate and cyanogen chloride are rarely formed under environmental conditions. These species are of interest primarily with respect to the alkaline chlorination process for treatment of cyanide (Chapter 20). Cyanate is unstable in water at neutral or low pH [3]. Cyanogen chloride is volatile and subject to degradation by hydrolysis in solution [3,4]. Organic forms of cyanide are compounds containing the –C≡N group. Common forms of organocyanides are the nitrile compounds, R−C≡N, where R is an alkyl group. Examples are methyl cyanide (CH3 CN, acetonitrile) and ethyl cyanide (CH3 CH2 CN). Cyanohydrins (2-acetoxy3-butenenitrile) and cyanogenic glycosides (amygdalin, dhurrin, and linamarin) are also part of the nitrile group. Figure 7.1 shows the molecular structure of several common organocyanide compounds. Figure 7.2 shows the chemical classification of selected dissolved inorganic cyanide complexes. As shown in the figure, the relative strength of the metal–cyanide bonds increases moving from left to right. Different groups of cyanide species are measured with different analytical methods. For example, the analytical methods for total cyanide capture the entire suite of cyanide compounds

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Increasing binding energy HCN Cd(CN)42– Zn(CN)42– Ag(CN)2– Cu(CN) 43– Ni(CN) 42– Hg(CN)o2 Au(CN)2– Pt(CN) 42– Fe(CN)4– Fe(CN)63– Co(CN) 63– 6

Free cyanide

Weak metal–cyanide complexes

Strong metal–cyanide complexes

Analytical definitions

CN amenable to chlorination; WAD CN

SAD CN (determined by difference)

Available CN

SAD CN (determined by difference)

Total CN

FIGURE 7.2 Chemical classification of dissolved inorganic cyanide species.

shown in Figure 7.2. On the other hand, the analytical methods for weak acid dissociable (WAD), cyanide amenable to chlorination (CATC), and available cyanide capture only those cyanide compounds with the weaker metal–cyanide bonds, that is, from mercury or nickel cyanide to free cyanide. Free cyanide is detected by all of the methods. The most common cyanide measurement performed is total cyanide analysis. This method involves addition of strong acid to a water sample followed by heating the water to boiling. These conditions lead to break up of metal–cyanide complexes, conversion of CN− to HCN, and volatilization of HCN(g). Air is bubbled through the boiling water to strip the HCN from solution. The HCN(g) distilled from the sample in this manner is captured in an aqueous sodium hydroxide solution through which the gas-flow from the distillation unit is passed. Analysis of the trap solution for cyanide ion is then performed to give the cyanide concentration in the sample. The next most common cyanide measurements performed on water samples are for CATC and WAD cyanide. Both of these techniques involve procedures similar to those employed in the total cyanide analysis technique. The reagents added to the water prior to heating and distillation are different from that in the case of the total cyanide analysis, however, for the purpose of liberating only the weakly complexed cyanide. The majority of the organocyanides are not measured with the conventional analytical tests, which are focused on inorganic cyanide species. Most organocyanide compounds are resistant to the release of cyanide ion under the conventional test conditions. There can be some partial recovery of organocyanide compounds in the conventional test methods [5]. Individual organocyanide compounds can be measured with liquid chromatography techniques, but there is no available method for measuring total organic cyanide in a water sample. Further information about the analytical techniques for organocyanide compounds is presented in Section 7.8. In this chapter, we describe the available analytical methods for cyanide species in water, and give some information about the performance of these methods with different kinds of waters. Table 7.2 lists analytical methods applicable for measurement of specific inorganic cyanide species or groups of species. Descriptions of each of the methods are presented, including target species and limitations. We begin with a discussion of interferences common to most of the methods, and pretreatments needed to avoid these interferences.

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SCN− CNO− CNCl

2− − 2− 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , Ni(CN)4 , 2− 4− 3− Hg(CN)2 , Hg(CN)4 , Fe(CN)6 , Fe(CN)6 , 3− − Pt(CN)2− 4 , Co(CN)6 , HCN, CN

2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , 2− 4− , , , Hg(CN) Hg(CN) Fe(CN) Ni(CN)2− 2 4 4 6 , 2− 3− − Fe(CN)3− , , , Pt(CN) Co(CN) HCN, CN 4 6 6

2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , − , HCN, CN Ni(CN)2− 4

− 2− Au(CN)− 2 , Ag(CN)2 , Ni(CN)4 , 4− 2− Fe(CN)6 , Cu(CN)4 , Co(CN)3− 6

2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , 2− 2− 3− Ni(CN)4 , Hg(CN)2 , Hg(CN)4 , Fe(CN)4− 6 , Fe(CN)6 , 2− 3− − Pt(CN)4 , Co(CN)6 , HCN, CN

2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , 2− 2− Ni(CN)4 , Hg(CN)2 , Hg(CN)4 , Fe(CN)4− 6 , 2− 3− − Fe(CN)3− , , , Pt(CN) Co(CN) HCN, CN 4 6 6

Thiocyanate Cyanate Cyanogen Chloride

Total cyanide

Total cyanide

Weak-acid dissociable cyanide

Weak acid and strong acid complexes

Total cyanide

Total cyanide

Available cyanide

Weak-acid dissociable cyanide Cyanide amenable to chlorination

2− − 2− 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , Ni(CN)4 , 2− − Hg(CN)2 , Hg(CN)4 , HCN, CN

Free cyanide

2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , 2− − Ni(CN)4 , HCN, CN 2− − 2− Cu(CN)2− 4 , Zn(CN)4 , Ag(CN)2 , Cd(CN)4 , 2− − − Ni(CN)4 , HCN, CN , SCN

Classification of species

CN− , HCN

Dissolved cyanide species Analytical method

Colorimetric Colorimetric Colorimetric

Automated using low-power UV digestion

Automated using UV digestion and thin film distillation

Automated using thin film distillation

Ion chromatography

Total cyanide after manual distillation

Flow injection ligand exchange

Flow injection ligand exchange

Weak-acid dissociable Amenable to chlorination

Microdiffusion

TABLE 7.2 List of Common Analytical Methods for Inorganic Cyanide Species

APHA 4500-CN M APHA 4500-CN L APHA 4500-CN J

USEPA 335.3 USGS I-2302

ASTM D4374

ASTM D4374

ASTM D6994 USEPA 9015

ASTM D2036 APHA 4500-CN C APHA 4500-CN D APHA 4500-CN E USEPA 9010B USEPA 335.2 USGS I-3300

OIA 1678

OIA 1677 (USEPA and ASTM)

ASTM D2036 APHA 4500-CN I ASTM D2036 APHA 4500-CN G USEPA 9010B USEPA 335.1

ASTM D4282

Method number and affiliation

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127

7.1 INTERFERENCES AND PRETREATMENTS Certain interferences and pretreatment methods are universal to most of the existing inorganic cyanide testing methods [3]. Cyanide samples are typically preserved with NaOH (pH ≥ 12) in order to avoid HCN volatilization losses during sample storage and to prevent precipitation of metal–cyanide complexes. Additionally, samples are cooled to 4◦ C to limit biodegradation of free cyanide. Some samples need to be filtered following preservation with NaOH. For example, samples with high iron content can form Fe(OH)3 (s) upon raising the pH. Filtration offers a convenient means of removing iron, which can form iron cyanide solids during distillation (blue distillate) causing a negative bias in total cyanide. This can be avoided by filtering the sample prior to distillation, in order to remove any colloidal or suspended iron oxides. Several chemicals can cause interferences with total cyanide measurements. In certain cases, it is advisable to remove these chemicals first, before raising the pH of the solution for sample preservation, to prevent interfering chemical reactions from occurring. The following are the common interferences in any kind of cyanide measurement. Oxidizing Agents. Oxidizing chemicals like chlorine can oxidize most cyanides during storage and handling. Spot testing and pretreatment for oxidizing agents should be performed prior to pH adjustment. Presence of such oxidizing agents should be tested by placing a drop of sample on a strip of potassium iodide (KI)–starch paper, previously moistened with acetate buffer at pH 4. Presence of any bluish coloration will indicate the presence of oxidizing agents. In that case, addition of 0.1 g of sodium arsenite per liter of sample is recommended to destroy the oxidizing agents [3], prior to sample preservation with NaOH. Reductants other than sodium arsenite, such as sodium thiosulfate and ascorbic acid, may also be used [3]. Sulfide. Sulfide and compounds of sulfide will convert CN− to SCN− rapidly at high pH [6]. Therefore, spot testing and pretreatment for sulfide should be performed prior to any pH adjustment. Lead acetate test paper moistened with acetate buffer (pH = 4) should be used to test samples for the presence of sulfide. If detected, pretreatment for sulfide includes addition of either lead acetate or lead carbonate, followed by filtration of precipitated lead sulfide. Many analytical laboratories have reported lack of sensitivity for the lead acetate paper spot test. An alternative practice is to pretreat the sample for sulfide without performing spot testing. In this case, 100 mg of either lead carbonate or lead acetate is added per liter of sample following acquisition. Following addition, the sample is filtered to ensure removal of lead sulfide and then immediately preserved with NaOH. Aldehydes. Aldehydes can convert any free cyanide ions to cyanohydrins, which can form nitriles during the distillation conditions, and nitriles are not measured by any conventional cyanide analytical methods. Various spot test reagents are available to detect the presence of aldehydes, for example, MBTH indicator solution, FeCl3 solution, and ethylenediamine solution. Further details about spot testing for aldehydes can be found in Standard Methods [3]. Pretreatment for aldehydes involves addition of 2 ml of 3.5% ethylenediamine solution to 100 ml of sample [3]. Fatty Acids. These organic acids can distill over to the alkaline absorber solution and can form soaps, thereby causing difficulty in detecting the endpoint during the titrimetric procedure. Fatty acids can be removed by extraction. Details are provided in Standard Methods [3]. Carbonates. At high concentration, carbonates can affect the distillation process by causing excessive generation of carbon dioxide (CO2 ) gas upon acid addition. CO2 (g) can also distill over to the absorber solution, thereby reducing the pH of the solution and resulting in inefficient capture of liberated HCN [7]. Whenever high carbonate conditions are suspected or detected (high alkalinity of the sample), calcium hydroxide should be used for pH preservation in place of sodium hydroxide. The calcium hydroxide is added to the sample slowly until the pH of the solution is between 12 and 12.5. Following precipitation of the carbonates, the sample is decanted for analysis. In addition to these chemical interferences, some of the metal–cyanide complexes are subject to photodecomposition, for example, iron-cyanide and cobalt-cyanide complexes, with the photolysis

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Aqueous sample

No change

No change

Chlorine check with KI–starch paper moistened with pH 4 acetate buffer

Sulfide check with lead acetate test paper previously moistened with acetate buffer (pH = 4)

Blue

Add 0.1 g Na2AsO2/l of sample

coloration

Darkening of the paper

Add 0.1 g PbCO3/l of sample

Filter sample

No

Presence of aldehyde suspected

Yes

Check with MTBH indicator

Yes

Add 2 ml 3.5% ethylenediamine to 100 ml of solution

No

Add NaOH raise pH between 12 and 12.5

No

Presence of fatty acid suspected

No

High carbonate content suspected (ALK > 500 mg/l)

Yes

Yes

Perform standard extraction with iso-octane/ hexane/ methylene chloride

Add Ca(OH)2 raise pH between 12 and 12.5

Perform cyanide analysis

FIGURE 7.3 Cyanide spot testing and pretreatment decision diagram.

rate faster at higher pH values [8–11]. Prior to pH preservation, prolonged exposure to sunlight can result in volatilization losses of photo-liberated HCN, thus reducing the concentration of total cyanide and metal–cyanide complexes. In pH-preserved samples, on the other hand, photodecomposition can result in a positive bias in free cyanide concentration. Care should be taken to minimize any light exposure during all procedural steps from sample collection to distillation and analysis. Figure 7.3 provides a decision diagram regarding spot testing and pretreatment. This chart provides guidance for when and how to pretreat samples prior to analysis.

7.2 TOTAL CYANIDE MEASUREMENT TECHNIQUES Operationally, the term “total cyanide” refers to all cyanide-containing groups that can be collectively measured as the cyanide ion, CN− , after some treatment steps to liberate the cyanide ion [3].

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Cooling water

129

In Out

To vacuum Condenser

Inlet tube

Caustic absorber

Distilling flask

Heater

FIGURE 7.4 Manual distillation apparatus for total cyanide test.

Total cyanide measurement can either be performed manually or by using automated techniques. As indicated in Figure 7.2 and Table 7.2, the total cyanide measurement does not capture all species bearing the −C≡N group. Further detail about the various total cyanide measurement techniques is provided below.

7.2.1 APHA/AWWA/WEF (STANDARD METHODS), ASTM, USEPA, AND USGS METHODS The most common total cyanide measurement is via manual distillation followed by colorimetric, titrimetric, or electrochemical finish techniques to measure cyanide ion concentration. The various analytical finish techniques are described in the following subsection. Figure 7.4 shows the total cyanide apparatus employed in the manual distillation procedure. The manual distillation technique is approved by all U.S. regulatory, government, and consensus organizations (see Table 7.2 for method numbers) and measures the free cyanide and metal-complexed (including both weak acid and strong acid dissociable) forms of inorganic cyanide present in an aqueous sample. The total cyanide method does not measure the following cyanide-related compounds: cyanates, thiocyanates, most organic-cyanide compounds, and most cobalt and platinum cyanide complexes. The detection limit in reagent water matrix is in the range of 1 to 5 µg/l. The conventional total cyanide method involves prolonged distillation of the sample at 125◦ C under strongly acidic conditions (pH < 2), which breaks apart most strong and weak

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metal–cyanide complexes and liberates free cyanide. Most organocyanides and some extremely strong metal–cyanide complexes such as cobalt cyanide do not fully decompose under the total cyanide distillation conditions. Additional sample predistillation treatment by ultraviolet radiation is required to obtain recovery of cobalt-cyanide complex. If incomplete recovery of cobalt cyanide occurs even after UV pretreatment followed by distillation, a second distillation is recommended [3]. The free cyanide present or generated by decomposition of complexes is converted to HCN under the acidic conditions and removed from solution by distillation as HCN(g). The distillate is passed through an alkaline absorber solution, which traps the HCN and converts it to the form of CN− . Final determination of CN− in the absorber solution is then performed via colorimetric, titrimetric, or cyanide ion-selective electrode methods. Standard Methods 4500-CN-C, Total Cyanide After Distillation [3]. Hydrogen cyanide (HCN) is liberated from an acidified sample (with sulfuric acid, H2 SO4 ) via a 2-h distillation and purging with air. HCN in the distillate is collected by passing the distillate gas through an NaOH scrubbing solution and then analyzed by any of the three finishing procedures, titration, colorimetric, or selective ion. The colorimetric analysis employs pyridine–barbituric acid for color development. This total cyanide method has been approved by USEPA for drinking water and NPDES compliance testing. ASTM Method D 2036-98, Standard Test Methods for Cyanides in Water, Method A — Total Cyanides After Distillation. This ASTM method [12] is technically similar to the Standard Methods total cyanide method 4500-CN-C except that it employs a 1-h reflux distillation compared to 2-h distillation for Standard Methods 4500-CN-C. Cyanide concentration can be quantified by either the titration, colorimetric, or selective ion electrode procedure. The method is approved by the USEPA for total cyanide determination. USGS Method I-4302-85, Total Cyanide, Colorimetric with Pyridine–Barbituric Acid. This method [13] is technically similar to Standard Methods 4500-CN-C and ASTM Method D 2036-98. USGS Method I-3300-85, Total Cyanide,Colorimetric with Pyridine–Pyrazolone. This method [14] is technically similar to ASTM Method D 2036-98 except that it utilizes pyridine–pyrazolone reagent for color development in the colorimetric procedure compared to pyridine–barbituric acid reagent used for ASTM Method D2036-98 and Standard Methods 4500-CN-C. This is also a USEPAapproved total cyanide analytical method. The above four methods, all use manual distillation with acidic reflux for the decomposition of complex cyanides prior to manual finish techniques for quantification of liberated cyanide. There are several semi-automated and automated analytical methods for the determination of total cyanide that are used primarily in municipal and industrial wastewater treatment plants. Some of the automated methods are presented below. USEPA Method 335.4, Determination of Total Cyanide by Semi-Automated Colorimetry. This USEPA method [15] utilizes a manual reflux-distillation operation to release hydrocyanic acid (HCN) from cyanide complexes, which is then absorbed in a scrubber containing sodium hydroxide solution. The cyanide ion in the absorbing solution is determined colorimetrically by automated, continuousflow analysis equipment designed to deliver and mix sample and reagents in the required order and ratios. USEPA Method 335.3, Total Cyanide by Colorimetric and Automated UV. This USEPA method [16] utilizes an automated UV digestion and distillation unit to decompose cyanide complexes and release HCN. Cyanides are determined automatically by a colorimeter and a recorder. This method was withdrawn for use in drinking water analysis by USEPA in 1994 [17] due to the concern about incomplete UV digestion, but still remains approved for reporting cyanide concentrations as required by NPDES permits. Total Cyanide by Low-Power UV Digestion Method. This is an automated analysis method implemented using an Skalar SAN plus segmented flow analyzer (model SA2001) with an SA1050 random access sampler, an SA 5570 in-line distillation unit, and an SA 555 UV-B inline digester

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(Skalar Analytical B.V., The Netherlands). The method is technically equivalent to the USEPA Method 335.3. Cyanide is released as HCN from cyanide complexes by means of UV digestion and distillation. HCN is then determined by a colorimetric procedure as described previously. ASTM Method D 4374-00, Standard Test Methods for Cyanides in Water, Automated Methods for Total Cyanide, Dissociable Cyanide, and Thiocyanate. This method [18] utilizes alkaline UV irradiation, acidification, and thin film distillation for cyanide-containing samples in an automated system. The breakdown of the strong metal–cyanide complexes, prior to the thin film distillation, is achieved by UV irradiation. Absorption of the liberated HCN gas is carried out using a glass coil and NaOH solution. It also employs the standard colorimetric determination of the recovered cyanides with an automated colorimeter. The method was developed and is employed routinely by the Municipal Water Reclamation District of Greater Chicago [19].

7.2.2 ANALYTICAL FINISH TECHNIQUES The three common analytical finish techniques used for final quantitation of cyanide ion liberated from the samples are titrimetric, colorimetric, and cyanide ion-selective electrode methods. Both the titrimetric and colorimetric procedures are approved by USEPA, Standard Methods (APHA / AWWA/ WEF), and ASTM; the cyanide ion-selective electrode method is approved only by Standard Methods. This section presents brief descriptions of each of the quantitation methods. Also discussed is amperometric measurement of cyanide ion, which is employed in the available cyanide by ligand displacement method. Titrimetric Procedure (Standard Methods 4500-CN-D, ASTM D2036-98). In this procedure [3,12], cyanide ion from the alkaline absorber solution following distillation is titrated with silver nitrate standard solution to form a soluble cyanide complex, Ag(CN)− 2 . Any presence of excess Ag+ in the solution is detected by a rhodanine indicator, which immediately turns from yellow to a salmon hue color indicating the endpoint. The concentration of cyanide in the absorber solution is calculated from the titrant normality and the volume of titrant used to reach the color endpoint. Typical calibration range is between 0.1 and 10 mg/l. Colorimetric Procedure (Standard Methods 4500-CN-E, ASTM D2036-98). In this procedure [3,12], the most commonly employed, cyanide ion from the alkaline absorber solution following distillation is converted to CNCl by reaction with chloramine-T at pH < 8. Following the formation of CNCl, pyridine–barbituric acid is added to the solution, which converts CNCl to a red–blue complex. The colored complex exhibits a stable absorption maximum at 578 nm. The concentration of cyanide in the absorber solution is determined spectrophotometrically by comparison against a standard calibration curve of absorbance vs. concentration. Typical calibration range is between 0.02 and 0.2 mg/l. Because of its detection sensitivity and broad calibration range, the colorimetric procedure is the analytical finish technique of choice for most cyanide analysis methods. Cyanide Ion-Selective Electrode (Standard Methods 4500-CN-F, ASTM D2036-98). The ionselective electrode is one of the few methods [3,12] that measures individual cyanide species, in this case CN− at high pH. In this procedure, cyanide ion from the alkaline absorber solution following distillation is determined potentiometrically using a CN− ion-selective electrode and associated meter. The concentration of cyanide ion in the absorber solution is determined by comparison against a standard calibration plot of log concentration of cyanide versus potential (mV). The cyanide electrode method is less sensitive in terms of detection than the colorimetric method and subject to numerous interferences. For these reasons, it is the least used of all the analytical finish techniques. Typical calibration range is between 0.05 and 10 mg/l. Amperometry (USEPA Method OIA-1677). In the “available cyanide by ligand displacement” method for measurement of free and weakly complexed cyanide [20,21], discussed in Section 7.3.2, liberated cyanide is isolated by use of a membrane, and the amount of cyanide ion collected is

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measured by amperometry. The cyanide isolation is achieved by acidifying the sample and allowing HCN to pass through a gas diffusion membrane. HCN is captured in a sodium hydroxide acceptor solution and thus converted to CN− ion. The use of a membrane selective for HCN makes it possible to employ a nonselective electrical conductivity measurement for quantitation of the amount of cyanide ion in the NaOH acceptor solution.

7.3 WEAK METAL–CYANIDE COMPLEXES There are three well-known analytical methods that render cumulative measurements of weak metal– cyanide complexes. These are (i) the weak acid dissociable cyanide (WAD) method; (ii) the available cyanide by ligand displacement method; and (iii) the cyanide amenable to chlorination (CATC) method. All of these analytical methods measure free cyanide in addition to the weak metal–cyanide complexes.

7.3.1 WEAK ACID DISSOCIABLE (WAD) CYANIDE METHOD The weak acid dissociable method is approved by ASTM [12] and Standard Methods [3] and involves distillation of the sample under slightly acidified (pH 4.5 to 6.0) conditions. This method does not recover CN− from strong metal–cyanide complexes, as indicated in Table 7.2. The most commonly used weak acid dissociable cyanide methods are Standard Methods 4500-CN-I “WAD Cyanide by Distillation” and ASTM Method D 2036-98, “Standard Test Methods for Cyanides in Water, Method C — Weak Acid Dissociable Cyanides.” In addition, there is an automated WAD cyanide method, ASTM Method D 4374-00, “Automated Method for Dissociable Cyanide by Thin Film Distillation [18].” Standard Methods 4500-CN-I, Weak Acid Dissociable Cyanide. Hydrogen cyanide (HCN) is liberated from a slightly acidified sample (acetate buffer, pH = 4.5) via a 2-h distillation and purging with air. HCN in the distillate is collected by passing the distillate gas through an NaOH scrubbing solution and then analyzed by any of the three finishing procedures, titration, colorimetric, or selective ion electrode (see Section 7.2.2 for analytical finish techniques). Figure 7.4 shows the instrument set-up for this method. The WAD cyanide method [3] is not yet approved by USEPA for drinking water or NPDES compliance testing, but has gained acceptance in several states (e.g., Pennsylvania and Texas). The WAD cyanide method has been observed to be less prone to interferences than the CATC Method [22,23] , which is a USEPA-approved method. The detection limit for this method is usually 1 to 5 µg/l for the colorimetric finish procedure. Zinc acetate buffer is used, prior to distillation, in the method to precipitate iron-cyanide complexes and enhance the selectivity of the method. However, for samples dominated by iron-cyanide complexes (>50%), an intermediate filtration step (using 0.45 µm filter) following zinc acetate buffer addition and prior to distillation is desirable to remove the precipitated zinc–iron-cyanide complexes. Otherwise, they can redissolve and dissociate under the conditions of the distillation and create a positive bias [24]. ASTM Method D 2036-98, Standard Test Methods for Cyanides in Water, Method C — Weak Acid Dissociable Cyanides. This ASTM method is technically similar to the Standard Methods weak acid dissociable cyanide method 4500-CN-I, except that it employs a 1-h reflux distillation in contrast to 2-h distillation for Standard Methods 4500-CN-I. Cyanide determination can be conducted colorimetrically, titrimetrically, or by cyanide ion-selective electrode. ASTM Method D 4374-00, Standard Test Methods for Cyanides in Water — Automated Methods for Total Cyanide, Dissociable Cyanide, and Thiocyanate. This is an automated method capable of detecting total cyanide, WAD cyanide, and thiocyanate [18]. For determination of WAD cyanide, the sample is acidified to pH 4.5, and exposed to a continuous thin film distillation unit to liberate HCN from the sample. Absorption of the liberated HCN gas is carried out using a glass coil and an NaOH

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Acceptor

Gas diffusion cell

Carrier

Acid

FIGURE 7.5 1999.)

Detector

Injection valve Pump

Waste

Mixing coil

Waste

Schematic for available cyanide Method OIA-1677. (Source: USEPA, Method OIA-1677,

trap solution. Cyanide determination is conducted colorimetrically at 578 nm by pyridine–barbituric acid reagent, as described in Section 7.2.2.

7.3.2 AVAILABLE CYANIDE BY LIGAND DISPLACEMENT METHOD USEPA and ASTM approved the available cyanide by ligand exchange method developed by OI Analytical (Wilsonville, OR), Method OIA-1677 [20,21,25], which measures mercury-cyanide complexes in addition to all the conventional weak acid dissociable complexes (i.e., cyanide complexes with Cu, Ni, Zn, Cd, Ag) and free cyanide. The list of analytes captured in the method is provided in Table 7.2. Researchers at the University of Nevada, Reno [26], in association with ALPKEM (a division of OI Analytical), developed this rapid, distillationless, flow-injection ligand-exchange method to determine available cyanide. The method consists of two parts: sample pretreatment, followed by cyanide quantification using amperometric detection. In the sample pretreatment step, ligand-exchange reagents are added to displace the cyanide ions from weak and intermediate strength metal–cyanide complexes. In the cyanide quantification step, a portion of the aliquot of the pretreated sample is injected into the flow injection manifold. The addition of hydrochloric acid converts cyanide ion to HCN, which diffuses through a membrane into an alkaline trap where it is reconverted to CN− . The alkaline trap solution is then analyzed for CN− amperometrically using a silver working electrode, a silver/silver chloride reference electrode, and a platinum counter electrode. Figure 7.5 shows a diagram of the flow injection system employed in the available cyanide method. The method detection limit in reagent water is approximately 0.5 µg/l. This method is particularly sensitive to any amount of sulfide in the solution, which is a positive interferent [23,27]. It is essential to treat samples with PbCO3 to remove sulfide prior to preservation when available cyanide analysis is to be performed. Section 7.1 should be consulted for information about the interferences and pretreatment steps required for any weak metal–cyanide complex analysis.

7.3.3 CYANIDE AMENABLE TO CHLORINATION This analysis method is applicable to those cyanide complexes and species that are “amenable to chlorination,” that is, upon chlorination, the cyanide complexes are dissociated and the liberated free cyanide is destroyed. In general, all weak metal–cyanide complexes (i.e., cyanide complexes of Cu, Ni, Zn, Ag, and Cd), are amenable to chlorination. Cyanides amenable to chlorination (CATC) are measured by separating and analyzing total cyanide in two sample portions: one portion chlorinated, and the other left as is and not chlorinated. Both sample portions are then subjected to the total cyanide analysis procedure. The difference in the CN− concentrations between the chlorinated and unchlorinated portions is designated as the cyanide amenable to chlorination. The CATC method has serious limitations owing to the fact that some organic compounds, including compounds that do not contain the −CN group as well as organocyanide compounds,

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can react to form free cyanide during chlorination, giving higher results for cyanide after chlorination than before chlorination [2,5,28,29]. This may lead to a negative value for the calculated concentration of cyanide amenable to chlorination, a problem that has often been encountered with steel industry samples, petroleum refinery distillate, and paper and pulp industry samples [3]. Also, the cyanide amenable to chlorination method has exhibited only partial recovery of some weak metal–cyanide complexes (e.g., nickel, silver, and mercury-cyanide complexes) in some matrices [26]. For the CATC method employing colorimetric finish, the detection limit of this method is usually 5 µg/l in reagent water matrix.

7.4 METAL–CYANIDE COMPLEX ANALYSIS BY LIQUID CHROMATOGRAPHY Various liquid chromatography techniques have been developed for the separation, identification, and quantification of the metal–cyanide complexes in water samples. Ion chromatography, which employs ion exchange resins for separation of metal–cyanide complexes, is the most common technique employed. Reversed-phase ion-pair partition chromatography, involving a nonpolar adsorbent, has also been used. Otu et al. [30] provide a review of the liquid chromatography techniques that have been developed for analysis of metal cyanides in water. A summary of the techniques is provided here.

7.4.1 ION EXCHANGE CHROMATOGRAPHY An ion chromatography method was developed by Dionex Corporation to measure specific metalcyanide complexes (i.e., cyanide complexes with Ag, Cu, Au, Ni, Fe, and Co) at mg/l levels of detection. The method may also be applicable for determining additional metal–cyanide complexes, such as platinum, and palladium cyanide complexes. This method has been approved by the USEPA [31] and by ASTM [32]. The original Dionex method has been modified to include both µg/l and mg/l level detection capability for the six target metal–cyanide complexes in that method [31,32], listed in Table 7.2. In aqueous solution, cyanide forms relatively stable anionic coordination complexes with most transition metals of the form [M(CN)x ]n− (M = the transition metal, x = the number of cyanide groups, and n = the electronic charge of the complex). Due to the stable nature of these complexes, they can be separated using anion exchange [31–33]. Following separation, detection is typically accomplished via low-wavelength ultraviolet light absorption at 215 nm. Figure 7.6 shows the ion chromatography instrumentation set-up for measurement of mg/l concentration levels of metal-cyanide complexes (high-level method). Determination of µg/l concentrations may also be accomplished using an automated online sample preconcentration [31,32]. Figure 7.7 presents typical chromatograms obtained for waters spiked with the six pertinent complexes. Method detection limits and calibration ranges for each metal–cyanide complex species are presented in Table 7.3. Further details about this method are available in ASTM [32]. The major interferences to this method are similar to those for the total and WAD cyanide methods, but there are some different considerations in analysis of metal–cyanide complexes by ion chromatography. Primary interferences and the corresponding pretreatments are discussed below. Photodecomposition. Some metal-cyanide complexes can photodissociate, resulting in decreases in their concentrations. Light exposure during sample collection, preparation, and analysis should be prevented, as much as possible. This applies to analytical samples as well as standards. Following acquisition, samples should be treated for any other interference (e.g., chlorine) and stored in amber bottles with preservatives. Chemical interferences. As in all ion chromatographic methods, certain chemicals can interfere with the analysis of metal–cyanide complexes and reduce the performance of the method, in terms of column performance and peak resolution. This becomes especially important when performing

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Anion trap column

Gradient pump

Autosampler Guard column

Separator column

Data analysis

UV-VIS detector

FIGURE 7.6 Ion chromatography instrument set-up for determination of metal–cyanide complexes. (Source: ASTM, Designation D 6994-04, Annual Book of ASTM Standards, Vol. 11.02, 2004. Copyright ASTM INTERNATIONAL. Reprinted with permission.)

[Au(CN)2]–

[Ag(CN)2]–

0.0

4.0

8.0

[Co(CN)6]3– [Cu(CN)3]2–

[Fe(CN)6]4–

12.0 Min.

16.0

Absorbance units

Absorbance units

[Cu(CN)3]2– [Ni(CN)4]2–

20.0

[Ag(CN)2]–

[Au(CN)2]–

0.0

4.0

8.0

[Co(CN)6]3– [Ni(CN)4]2– [Fe(CN)6]4–

12.0 Min.

16.0

20.0

FIGURE 7.7 Ion chromatograms from analysis of metal–cyanide complexes in reagent water (left) and groundwater sample (right). Metal–cyanide species concentrations are in the mg/l range. (Source: ASTM, Designation D 6994-04, Annual Book of ASTM Standards, Vol. 11.02, 2004. Copyright ASTM INTERNATIONAL. Reprinted with permission.)

sample preconcentration. Carbonates, organic acids, and high total dissolved solids can contribute to unstable baselines and large front-end tailing. While the presence of such species may not be avoided in samples, the use of high purity water for reagent and eluent preparation is essential for ensuring the highest quality chromatography. As most samples for cyanide analysis are preserved at pH 12 or greater, optimum results are achieved when the matrix of the calibration standards is matched to those of the samples.

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TABLE 7.3 Method Detection Limits and Calibration Ranges for the Analysis of Individual Metal– Cyanide Complexes by Ion Chromatography Study level

Ag(CN)− 2

Au(CN)− 2

Co(CN)3− 6

Cu(CN)2− 3

Fe(CN)4− 6

Ni(CN)2− 4

High level (mg/l)

MDL Calibration range

0.77 1–100

0.64 1–50

0.43 1–100

0.09 0.1–2

0.09 0.1–20

0.83 1–200

Low level (µg/l)

MDL Calibration range

8.66 10–125

2.8 5–100

0.99 1–200

4.56 0.1–5

0.21 0.5–20

7.33 50–100

Source: From ASTM, Designation D 6994-04. Annual book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2004.

Oxidizing agents. Oxidizing agents like chlorine can decompose certain weak metal-cyanide complexes, thereby causing a decrease in their concentration. Refer to Section 7.1 for pretreatment steps.

7.4.2 REVERSED-PHASE ION-PAIR PARTITION CHROMATOGRAPHY In ion pair (or ion interaction) chromatography, ion pairs are partitioned between a polar mobile phase and a hydrophobic stationary phase. This approach has been applied for the chromatographic separation and quantification of metal–cyanide species in water samples [34–37]. There is no standard method that has been adopted by governmental or consensus organizations, as the technique has been applied mostly in research contexts. However, Waters Corporation has published a method [38] based on the approach of Hilton and Haddad [36]. The ion-pair chromatography methods that have been developed are quite similar to one another. In the methods used by Hilton and Haddad [36] and Grigorova et al. [34], for example, a C-18 stationary phase is employed (C18 Novapak cartridge column, Waters Corporation), and the mobile phase consists of a solution of 2 to 5 mM tetrabutylammonium hydrogensulfate-methanol (approximately 70:30 by volume). Detection is by UV absorption. The mobile phase reagents cause minimal background interference in the low-UV range used in the UV detector, for example, 205 to 215 nm. Elution and separation of metal–cyanide complexes progress in order of decreasing ion-pair polarity, with the more polar ion pairs being eluted earlier. Experience with ion-pair chromatography indicates that the method is capable of measuring rapidly a range of metal–cyanide complexes at mg/l levels. Grigorova et al. [34], for example, − 4− 3− demonstrated the ability of the method to distinguish Cu(CN)− 2 , Ag(CN)2 , Fe(CN)6 , Co(CN)6 , 3− − Ni(CN)2− 4 , Fe(CN)6 , and Au(CN)2 when present together at 10 mg/l each in a synthetic solution. Separation and measurement were completed in 30 min. Experiments with hydrometallurgical gold mining effluent yielded similar results, with individual metal–cyanide species separated, identified, and measured rapidly at mg/l levels. The limited efforts made to apply ion-pair chromatography for measurement of metal–cyanide species at lower (µg/l) level concentrations, indicate the need for further method development. Haddad and Kalambaheti [39], using an analytical approach similar to that described above, studied recovery of various metal-cyanide complexes at low concentrations. A preconcentration step was implemented in the method. The studies were conducted with various electrolyte compositions. 3− It was found that recovery from spiked solutions was low for Fe(CN)4− 6 and Cu(CN)4 , and significant interferences were caused by the presence of other common anions such as Cl− and SO2− 4 .

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Overall, available information indicates that the ion-pair chromatography method is promising, but needs further development for measurement of metal–cyanide species in complex waters at µg/l concentrations. Haddad and Kalambaheti [39] identified some potential ways to address the problems they encountered. There apparently has been little effort to develop the method further, however, since the work in the late 1980s to early 1990s. Huang et al. [37] worked to optimize various aspects of the method, but their efforts were not focused on enhancing method performance at low concentration levels.

7.5 FREE CYANIDE ANALYSIS In analyzing the cyanide content of a water sample, free cyanide, that is, the sum of HCN and CN− , is usually of primary interest, as HCN is a highly toxic form of cyanide. The cyanide anion CN− is easily converted to HCN via a simple change in pH (pK a for HCN is approximately 9.2 at 25◦ C). Since the 1950s, there has been sustained interest and effort in development of reliable, rapid, lowcost techniques for measuring free cyanide at low (µg/l) concentrations in water samples of varying composition. These efforts have yielded a number of different methods, only a few of which are in widespread use today. The differentiating factors among these methods are primarily cost and ease of application. In this section, the leading approaches for analysis of free cyanide in water are described. The most widely used methods are identified, and the reasons for their widespread adoption are discussed.

7.5.1 GAS CHROMATOGRAPHY Since HCN is a volatile species, gas chromatography (GC) may be used for measurement of HCN content in water. GC techniques for HCN analysis were initially developed in the 1960s [40,41] and used in the performance of the first definitive studies of cyanide toxicity to aquatic organisms [42]. In developing the GC technique for HCN analysis, an objective of Schneider and Freund [41] was to avoid shifting the HCN/CN− equilibrium. The technique of Schneider and Freund [41], later modified by Claeys and Freund [40], involves first stripping HCN from an unaltered water sample by passing finely dispersed compressed air through the sample. A relatively large volume sample is used (e.g., 20 l) so that the stripped HCN comprises only a small portion of the HCN in the system. The gas exiting the sample is then passed through a heated drying tube, and into a cooled concentration column containing an adsorbent coated on a granular support. The HCN is thus cold trapped in the concentration column. Subsequently, the concentration column is heated and the contents of the column are injected via carrier gas into a gas chromatograph equipped with a thermal conductivity detector [41] or flame ionization detector [40]. Claeys and Freund [40] demonstrated that HCN concentrations as low as 1 µg/l can be detected with this technique. Modern gas chromatography equipment permits measurement of HCN at sub-ppb concentrations. While purge and trap GC techniques for measurement of HCN are well established, the approach is not widely used, primarily because of the expense of the analysis. Unlike the case for volatile organic compounds, for which GC is used extensively, the instrument configuration needed for HCN analysis often is not useful for the simultaneous measurement of other compounds of interest. Less expensive techniques with detection limits and accuracy similar to that of gas chromatography are available, and are used more frequently.

7.5.2 DIRECT COLORIMETRIC DEVELOPMENT The colorimetric procedure used to measure the amount of cyanide ion present in the NaOH scrubber solution for the total cyanide test (Section 7.2.2) is sometimes used directly on water samples, without distillation. For example, in many of the studies used as the basis for U.S. ambient water quality

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criteria for cyanide [43], free cyanide was determined using the pyridine–pyrazalone colorimetric method from the 13th edition of Standard Methods [44]. In that earlier edition, the introduction to the total cyanide method indicates that distillation can be omitted if “it is known that the sample contains only simple cyanides of the alkalis and is completely free of all interferences [44].” The same document has a list of interferences that includes sulfides, heavy metal ions, fatty acids, thiocyanate, cyanate, glycine, urea, and oxidizing agents. The possibility of omitting the distillation step in total cyanide analysis is not mentioned in more recent editions of Standard Methods [3]. In fact, this is now explicitly discouraged: “The importance of the distillation procedure cannot be overemphasized [3].” Direct colorimetric development to analyze free cyanide is not a standard method of any U.S. government agency or consensus organization. While it is still practiced by some, there are important calibration and interference issues that make the accuracy of the method doubtful, especially with higher ionic strength water samples. With respect to calibration, for example, the issue of matrix matching is significant. Calibration standards for colorimetric analysis of the distillate in the total cyanide test are prepared in NaOH solution of the same composition as the NaOH scrubber solution in which the HCN in the distillate gas is trapped. Using a calibration curve based on standards in NaOH solution to interpret color developed directly in a water sample is problematical. The chemical composition of the calibration standard solution should match that of the sample. Another source of error in direct colorimetric development is pH variation between sample and calibration standards. Variations in pH will affect the cyanide ion distribution in the sample. Calibration standards in NaOH solution are all at a common pH. Finally, for samples dominated by metal–cyanide complexes, it will be difficult to quantify the free cyanide content of the sample with a reasonable degree of accuracy, without first separating the complexes from the solutions to be analyzed for free cyanide. All of the problems listed are eliminated by inclusion of distillation prior to colorimetric development. The accuracy of direct colorimetric development for measurement of free cyanide was evaluated by Dzombak and Higgins [45]. To examine the performance of the direct colorimetric development technique for measurement of free cyanide, spiked samples of freshwater and seawater were tested using the 1971 Standard Methods direct colorimetric method [44], the current colorimetric procedure [3], and the free cyanide by microdiffusion method [46]. The 1971 direct colorimetric method [44] is also sometimes referred to as the “pyridine–pyrazolone method.” The current colorimetric procedure [3] uses pyridine–barbituric acid for color development. Samples of filtered water from a moderately hard, freshwater lake in Pennsylvania and filtered seawater from coastal Oregon were spiked with 20 to 150 ppb free cyanide (KCN) and then analyzed using the methods cited. Results are presented in Table 7.4. It was found that direct colorimetric analysis for measuring free cyanide yielded results comparable to, though less consistent than, those obtained by microdiffusion for free cyanide in freshwater, but that direct colorimetric analysis was grossly inaccurate for measuring free cyanide in seawater. The clear failure of direct colorimetric analysis with the seawater was most likely due to the high ionic strength of the seawater, a known interferent for the colorimetric procedure, and difficulty in buffering the analysis solution to the correct pH.

7.5.3 GAS–LIQUID DIFFUSION Various techniques have been developed that exploit the volatility of HCN, to measure free cyanide in aqueous samples by allowing the HCN to volatilize and then by capturing and measuring the trapped HCN. The basic approach involved with these methods is to provide a confined gas (air) volume above a water sample, and to also put in contact with the gas phase an NaOH solution that serves as a sink for the HCN that diffuses out of the water sample and into the gas phase. Kruse and Thibault [47] first proposed the use of a microdiffusion cell for measurement of free cyanide in water samples. A Conway microdiffusion cell (Figure 7.8) consists of two concentric

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TABLE 7.4 Resultsa of Free Cyanide Analyses in KCN-Spiked Freshwater and Seawater Samples using Direct Colorimetric Analysisb and Microdiffusionc [CN− ] using direct colorimetric analysisb (µg/l)

[CN− ] using free cyanide by microdiffusionc (µg/l)

Percent of free (diffusible) cyanide detected using colorimetric method

Freshwater 20 60 150

36.9 69.2 144

20.2 58.2 145

183 119 99.7

Seawater 20 60 150

11 1.9 4.8

20.4 58.5 155

53 3.3 3.1

Sample nominal value (µg/l)

a Source: Data from Dzombak, D.A. and Higgins, C.J., Quarterly Progress Report (May 27)

WERF Project 01-ECO-1, Water, Environment Research Federation, Alexandria, VA, 2004. b APHA Method 207, Standard Methods for the Examination of Water and Wastewater,

13th ed., American Public Health Assoc., American Water Works Assoc., and Water Environment Research Federation, Washington, DC, 1971. c ASTM D 4282-95. Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998.

NaOH absorber solution

Sample

FIGURE 7.8 Test apparatus for analysis of free cyanide by microdiffusion.

chambers and a plastic cover or lid. The water sample is introduced in the outer chamber, while NaOH absorber solution is introduced in the inner chamber. After addition of the sample to the outer chamber, a pH 6 buffer solution is added to the sample and the cover or lid is replaced. The entire cell is then placed in the dark for a specified period of time, to allow passive diffusion of HCN to occur. At the end of the diffusion period, the NaOH absorber solution is analyzed for free cyanide content by the colorimetric method outlined in Section 7.2.2. From tests with spiked solutions at different pH conditions, Kruse and Thibault [47] found that complete recovery of free cyanide could be obtained at pH 7 in diffusion periods of 5 h or less.

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The microdiffusion method subsequently was recommended by the American National Standards Institute for measuring free cyanide in photographic development process effluents [48]. The method gained more widespread use and was added as a standard method for free cyanide analysis by ASTM in 1983 [46], with a modification that involved the addition of CdCl2 as a precipitation agent for iron-cyanide complexes following the addition of pH buffer to the sample in the outer chamber of the microdiffusion cell. Broderius [49] developed a similar diffusion procedure for isolating and concentrating HCN in NaOH, but he employed a larger reaction vessel (flask) containing 1.5 l of test solution and about 1.3 l of gas phase volume. A glass dish containing 8 to 20 ml of 0.02 N NaOH was suspended above the test solution inside the flask. A 2 h diffusion period was employed in HCN testing with the apparatus. Broderius chose a larger-scale reactor design and relatively short diffusion period in order that only a small fraction of the total HCN would be removed during the test, thus not disturbing significantly the cyanide species equilibrium distribution in the sample. Also, the Broderius [49] test procedure involved no reagent additions to the water sample for pH adjustment and control. Broderius explicitly wanted to avoid sample acidification as he expected this to induce “conversion to HCN of simple cyanides and, most likely, some portion of certain metallocyanides.” While the Broderius procedure is advantageous for assessing the true HCN concentration under particular solution conditions, it is more often of interest to know the total free cyanide concentration (HCN + CN− ) in a sample. Free cyanide is defined as the sum of the two species because CN− is converted instantaneously to HCN upon a simple lowering of pH in a solution that contains it. Moreover, except for very high pH samples and samples with significant amounts of weak metal– cyanide complexes present, microdiffusion and the Broderius method will yield the same results as HCN is the dominant form of free cyanide for all systems with pH < 9.2. For these reasons, and because of the small scale and simplicity of the method, the microdiffusion method has become the standard and preferred gas–liquid diffusion method for measuring free cyanide in water. Free cyanide by microdiffusion is an ASTM approved method [46] and involves the measurement of HCN evolved passively from an acidified sample. It is a simple technique and easily applied to waters of complex composition. With respect to the microdiffusion method, free cyanide is defined as the cyanide that diffuses at room temperature from simple cyanides or weak metal–cyanide complexes as hydrogen cyanide gas, from a solution of pH 6 to 6.5. The test method does not measure metal– cyanide complexes or organocyanide compounds that resist dissociation, such as, iron cyanide, cyanohydrin, etc. The microdiffusion method does recover some fraction of weaker metal–cyanide complexes if present [23]. The ASTM microdiffusion method is performed in a Conway microdiffusion cell (Figure 7.8). The water sample is introduced in the outer chamber, while a specified amount of NaOH absorber solution is introduced in the inner chamber. After the addition of the sample to the outer chamber followed by the addition of cadmium chloride solution to precipitate the hexacyanoferrates, a pH 6 buffer solution is added to the sample and the cover or lid is replaced. The closed cell is then placed in the dark to allow 4 h of diffusion [46]. During this time, free cyanide diffuses as HCN gas and is absorbed into the sodium hydroxide absorber solution in the center chamber, where it is converted to CN− . At the end of the specified diffusion period, the CN− in the absorber solution is analyzed using the standard colorimetric procedure (Section 7.2.2). The typical calibration range is between 0 and 150µg/l. In addition to the potential for weak metal–cyanide complexes to decompose under the acidic conditions imposed for the microdiffusion test, the method is subject to some other interferences, primarily from other volatile species that may also be trapped in the NaOH absorber solution. Volatile species that can affect the results obtained in the microdiffusion test include ammonia, sulfide, and phenol [50]. These substances need to be present in fairly large quantities in the NaOH absorber solution, in order to influence the colorimetric procedure used to evaluate its free cyanide content, for example, ammonia and phenol greater than 50 mg/l, and sulfide greater than 1 mg/l. The microdiffusion method has undergone an inter-laboratory study and has been modified to eliminate the use of CdCl2 for iron-cyanide complex precipitation [51]. The revised method will

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be submitted to the USEPA Office of Solid Waste for potential inclusion in the USEPA analytical methods manual SW-846 [52].

7.5.4 ION-SELECTIVE ELECTRODE As discussed in Section 7.2.2, the cyanide ion-selective electrode can be used for measurement of CN− concentration in the NaOH absorber solution used to capture HCN from the distillate in conventional total cyanide analysis. The cyanide ion-selective electrode is also employed directly for analysis of free cyanide in water samples. For example, in the United States, the cyanide ion-selective electrode is allowed to be used directly for analysis of finished drinking water, to demonstrate compliance with the U.S. national primary drinking water regulations [53]. The cited procedure, Method 4500CN F from Standard Methods, is presented there as a finish procedure to analyze CN− ion in the NaOH absorber solution used to trap HCN in the distillation of a water sample for total cyanide analysis. Use of the cyanide ion-selective electrode for direct analysis of a water sample would require dilution with NaOH prior to analysis, to match the matrix of calibration standards and to convert all HCN to CN− . There is no discussion in Standard Methods [3] of the use of the ion-selective electrode directly on a water sample made alkaline by NaOH addition, but this approach has been employed and reported in the literature. Sekerka and Lechner [54] presented a method for analysis of free cyanide in a water sample using the cyanide ion-selective electrode without distillation. To a 20 ml sample of water, 10 M NaOH is added to achieve a pH of 11.5. Small quantities of some additional reagents (orthophosphoric acid and bismuth nitrate), related to a procedure for identification of interfering halides, are also added. Cyanide ion concentration is determined by the standard colorimetric method. In tests with analysis of synthetic solutions of KCN in double-distilled water, cyanide ion concentrations as low as 10−7 M(2.6 µg/l) were measured. Sekerka and Lechner [54] also studied the magnitude of known interferences with the cyanide ion-selective electrode. The cyanide electrode malfunctions if anions that form salts with silver are present, including sulfide (S2− ), iodide (I− ), bromide (Br − ), thiocyanate (SCN− ), and chloride (Cl− ). Measurements performed on samples with 5 × 10−7 M cyanide ion indicated that chloride and thiocyanate did not interfere until their concentrations exceeded the cyanide concentration by 104 , and bromide did not interfere up to a concentration ratio of 102 . On the other hand, iodide ion interfered at a 1:1 ratio, and sulfide species interference was significant at all concentrations studied. It is thus feasible to employ the cyanide ion-selective electrode for direct analysis of free cyanide in water, but only for relatively simple, low ionic strength water samples. Without distillation, there is a large potential for interference with the cyanide ion-selective electrode measurement from ions and organic substances present in the water. Because of this, and the availability of less complicated alternative methods such as the microdiffusion technique [46], the cyanide ion-selective electrode is infrequently used for direct measurement of free cyanide on water samples without distillation.

7.6 THIOCYANATE AND CYANATE MEASUREMENT TECHNIQUES 7.6.1 THIOCYANATE MEASUREMENT PROCEDURE Thiocyanate (SCN− ), which is not captured by conventional total cyanide tests, may be measured colorimetrically using Standard Methods 4500-CN M [3]. The procedure involves sample titration under acidic pH conditions (pH = 2) with ferric nitrate. Under the analysis conditions, the Fe3+ reacts with SCN− to form an intense red color complex that exhibits an absorption maximum near 460 nm. The concentration of thiocyanate is determined spectrophotometrically at the wavelength maximum by comparison against a standard calibration curve of absorbance versus concentration.The

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typical calibration range is between 0.1 and 2 mg/l. Further details about the method are available in Standard Methods [3]. Common interferences to the method and their corresponding remedies are discussed as follows: Hexavalent Chromium. Hexavalent chromium is a positive interference and is removed by addition of ferrous sulfate after acidifying the sample to between pH 1 and 2 using nitric acid. Reducing Agents. Reducing agents can reduce Fe3+ to Fe2+ , thereby preventing the formation of the color complex. Reducing agents can be destroyed by adding few drops of hydrogen peroxide. Sulfide. Sulfide can convert any free cyanide ion in the sample into SCN− during the test procedure and therefore could serve as a positive interferent. Sulfides are removed by adding lead salts prior to the iron addition. Refer to Section 7.1 for sulfide removal steps. Microbes. Thiocyanate is biodegradable. Samples are thus preserved with acid addition to pH < 2 and stored at 4◦ C to inhibit microbial activity. In addition to the titration method, an ion chromatography technique is also available to measure SCN− in solutions. Separation of thiocyanate is achieved using an AG5 guard column and an AS5 analytical column with AMMS-I and 25 mM H2 SO4 as column suppressor and regenerant, respectively [33]. Isocratic elution is maintained using the four different eluent preparations (2 mM NaOH, 4.5 mM Na2 CO3 ; 2 mM NaOH; 0.8 mM 4-cyanophenol and 2% acetonitrile) and final detection is performed using a conductivity detector. This method has not yet been approved by any consensus or governmental organization.

7.6.2 CYANATE MEASUREMENT PROCEDURE Cyanate is also not captured in conventional total cyanide tests and must be measured individually when its presence is suspected and is of interest. The Standard Methods 4500-CN-L [3] procedure for measuring cyanate (CNO− ) involves high temperature hydrolysis at low pH. At a pH of 2 to 2.5 and at a temperature between 90 and 95◦ C, cyanate hydrolyzes to ammonia according to the following reaction: 2NaCNO + H2 SO4 + 4H2 O → (NH4 )2 SO4 + 2NaHCO3

(7.1)

Ammonia concentration is measured before and after the treatment using either an ammonia ion-selective electrode, or by colorimetric development followed by direct nesslerization. This test is applicable for measurement of cyanate compounds in industrial waste and natural waters. The typical calibration range is between 1 and 200 mg/l with method detection limit around 1 to 2 mg/l. Common interferences to the method and corresponding pretreatments are as follows: Organic Compounds. Organic nitrogenous compounds can hydrolyze to ammonia following acidification. Thus, acidification and heating must be controlled carefully to minimize such interferences. Metals. Metallic compounds can form colored complexes with nessler reagents during ammonia analysis. Addition of Rochelle salt or EDTA during the ammonia analysis by nesslerization can overcome these interferences. Oxidizing agents. Certain powerful oxidizing agents can oxidize cyanate to CO2 and nitrogen. Refer to Section 7.1 for detection and pretreatments of oxidizing agents. Further details about the method are available in Standard Methods [3].

7.7 CYANOGEN HALIDE ANALYSIS Cyanogen halides, especially cyanogen chloride (CNCl), are highly toxic species and are not captured in the total cyanide test. Hence, their formation and presence is of interest, especially in situations in which water containing cyanide is chlorinated or brominated [4,29].

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Cyanogen halides (CNX, where X = Cl, Br, etc.) react with pyridine barbituric acid (see Section 7.2.2 under Colorimetric Procedure) to produce a red–blue color complex at pH < 8 that exhibits an absorption maximum at 578 nm. The concentration of cyanogen halides is determined spectrophotometrically by comparison against a standard calibration curve of absorbance versus concentration. The typical calibration range is between 5 and 150 µg/l. A colorimetric technique specifically for cyanogen chloride is given as Method 4500-CN J in Standard Methods [3]. The primary interference to this method is the instability of the analytes [3]. Cyanogen halides can hydrolyze to cyanate at a pH of 12 or more. Standard Methods [3] recommends that samples should be collected without any form of preservation, including NaOH addition, and analyzed as quickly as possible following acquisition to minimize any hydrolysis. If the presence of oxidizing agents or sulfide is detected or suspected, pretreatment steps outlined in Section 7.1 should be implemented. Gas chromatography can also be used for analysis of cyanogen halides [4,55]. This approach has been used primarily for analysis of drinking water samples. To prevent the hydrolysis degradation of cyanogen halides during sample storage, it is recommended that the pH of samples be lowered to 3.0–3.5 with sulfuric acid [4].

7.8 ORGANOCYANIDE MEASUREMENT TECHNIQUES Organocyanide compounds generally are not measured with the conventional analytical tests, which involve treatment to liberate free cyanide followed by measurement of the free cyanide. Most organocyanides are resistant to the release of cyanide ion in the total cyanide, WAD cyanide, and CATC test conditions. There can be some partial recovery of organocyanide compounds in the conventional test methods. Table 7.5 presents results obtained by Yi et al. [5] for recovery of cyanide from solutions spiked with four different organocyanide compounds in the standard total cyanide test. As shown there, for two of the compounds there was essentially no recovery, while for the other two compounds, mass recoveries (as CN) were 19 and 73%, respectively. The effect of chlorinating the solutions on recovery was also studied. There was no change in recovery for the two compounds not detected in the total cyanide test without chlorination (Table 7.5). For the other two compounds, recovery was reduced to zero for one and doubled for the other. The reduction in recovery probably resulted from oxidative destruction of the compound, including the −CN group. Overall, the results of Yi et al. [5] demonstrate that recovery of cyanide from organocyanide compounds in the standard total cyanide and CATC method is quite variable and substantially less than 100%. Methods other than the conventional cyanide analysis techniques need to be employed to analyze for organocyanide compounds in water. Specific organocyanide compounds can and have been measured with liquid chromatography techniques [56], but there is no available method for measuring total organic cyanide in a water sample. Theis et al. [57] developed and tested a scheme for assessment of total organocyanide content of a water sample that involved the following steps: • Separation of inorganic cyanides by passing the preserved sample (pH ≈ 11) through an anion exchange resin bed (100 to 200 mesh resin). • Measurement of concentrations of inorganic species remaining after ion exchange pretreatment (<10 ppb) by microdiffusion and ion exchange chromatography. This quantifies inorganic cyanide content remaining after pretreatment, for subtraction from the apparent organic cyanide content measured in the last step. • Ozonation of the pretreated sample (with pH ≈ 11) to oxidize the cyanide group (−C≡N) in the organocyanides to cyanate (CNO− ) and then nitrate (NO− 3 ), which are analyzed by ion chromatography to provide an indirect measurement of the total organic cyanide

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TABLE 7.5 Percentage Recovery of Cyanide from Four Organocyanide Compounds in Total Cyanide and Cyanide-Amenable-toChlorination Testsa Organocyanide compoundb Acetonitrile d-Amygdalin 2-Acetoxy-3-butenenitrile Cyanocobalamin

CN recovery (%) in total cyanide testc

CN recovery (%) in CATC testd

<1.57 <1.38 73.40 ± 5.91 18.98 ± 7.67

<1.57 <1.38 <1.46 41.06 ± 6.11

a Source: Data from Yi, Y., Thesis, T.L., and Young, T.C., Water Environ. Res.,

74, 51, 2002. Copyright 2002 Water Environment Federation, Alexandria, Virginia. Reprinted with permission. b Conc. = 2 ppm. c USEPA Method 335.2. Rev. 1980, Methods for the Chemical Analysis of Water and Wastes EPA-600/4-79-020, U.S. Environmental Protection Agency, Cincinnati, OH, 1979. d USEPA Method 335.1. Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, Cincinnati, OH, 1979.

present. The reaction path involved in ozonation of free cyanide was demonstrated in preceding work [58]. In testing this approach with synthetic solutions of inorganic and organic cyanide compounds (2 to 5 mg/l as CN), it was found that inorganic cyanide species could be removed very effectively by anion exchange, and that organocyanide compounds could, indeed, be converted to cyanate and nitrate. However, it was also determined that rates of oxidative conversion of the organocyanide compounds during ozonation varied widely. It was concluded that complete oxidative conversion was not feasible within the practical time of ozonation in an analytical test of a few hours. In addition, it was determined that thiocyanate (SCN− ) was converted to cyanate and nitrate very rapidly, indicating that it could pose an interference problem with the ozonation-based technique. For these reasons, the proposed total organic cyanide technique was not pursued further. Considering the approach of Theis et al. [57] has value, however, for appreciation of the challenges involved in developing a total organic cyanide technique, and for guidance to others, who will undertake this challenge in the future.

7.9 COMPARATIVE METHOD PERFORMANCE The methods in use for measurement of individual and groups of cyanide species have capabilities and limitations that depend on water composition. As documented earlier in this chapter, there are a number of potential interfering agents that are common to many of the cyanide analytical techniques. Even with pretreatment for these common interferences, certain methods will still have difficulty in waters of complex composition. Zheng et al. [23] evaluated the comparative performance of seven methods of cyanide analysis in measuring different cyanide species in reagent water and five different contaminated water

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TABLE 7.6 Contaminated Water Matrices Evaluated in Cyanide Methods Performance Study by Zheng et al.[23] ID no.

Site/plant name

Site/plant type

Sample type

CW1

Manufactured gas plant site groundwater, NY

Former MGP site

CW2

Hanover Park Water Reclamation Plant, Metropolitan Water Reclamation District, Hanover Park, IL Hanover Park Water Reclamation Plant, Metropolitan Water Reclamation District, Hanover Park, IL Deer Island Treatment Plant, Massachusetts Water Resources Authority, Boston, MA Aluminum smelting plant, TN

Publicly-owned treatment works

Contaminated groundwater Unchlorinated POTW secondary effluent

Publicly-owned treatment works

Chlorinated POTW secondary effluent

Publicly-owned treatment works Waste disposal site

POTW primary clarifier effluent Contaminated groundwater

CW3

CW4 CW5

Source: Reprinted with permission from Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Environ. Sci. Technol., 37, 107, 2003. Copyright 2003 American Chemical Society.

matrices. The seven methods evaluated included five species-specific methods — weak acid dissociable cyanide, free cyanide by microdiffusion, available cyanide, automated WAD cyanide by thin film distillation, metal cyanides by ion chromatography — and two automated techniques for total cyanide — total cyanide by thin film distillation, and total cyanide by low-power UV digestion. Total cyanide was also measured in each contaminated water sample by the conventional distillation technique, but this widely used and well known technique was not part of the methods comparison experiments. The goal of the study was to assess the detection limits, accuracy, and precision of the seven alternative cyanide analysis methods with a broad spectrum of contaminated water matrices. Emphasis in the study was on the performance of the methods at low (<100 ppb) cyanide species concentrations in municipal and industrial water matrices. The five contaminated water matrices employed by Zheng et al. [23] are listed in Table 7.6. As seen there, the water samples collected for the study included publicly owned treatment works (POTW) influents, POTW effluents, and contaminated groundwaters from a former manufactured gas plant (MGP) site, and from a waste disposal site at an aluminum smelting plant. Characterization data for the five contaminated waters used in the study are provided in Zheng et al. [23]. Samples of the contaminated waters as well as clean reagent water were spiked with NaCN, K2 Ni(CN)4 , and K4 Fe(CN)6 at levels up to 150 µg/l. For each combination of contaminated water, cyanide species spiked, and spike amount, numerous replications were performed in accordance with procedures specified by the USEPA and Standard Methods for quantification of method detection limit, bias, precision, and mass recovery [23]. The method detection limits (MDLs) determined from the analysis of replicate spiked samples are presented in Table 7.7. As indicated by the results shown in Table 7.7, almost all of the analytical methods were capable of MDLs at low ppb levels for their target analytes or analyte groups. The MDLs obtained were less than 5 ppb in most cases. Other results obtained by Zheng et al. [23] indicated that the cyanide analytical methods evaluated exhibited satisfactory accuracy and precision for most of the contaminated waters used in the study. Problems occurred with analysis of low concentrations of cyanide species in raw municipal wastewater (CW4) for the available cyanide and ion chromatography methods, which experienced

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TABLE 7.7 Method Detection Limits for Various Cyanide Analytical Techniques in Contaminated Water Matrices Method detection limits (ppb) Method

Water matrix

NaCN

K2 Ni(CN)4

K4 Fe(CN)6

Available CN (EPA OIA-1677)

DIWa CW1 CW2 CW3 CW4 CW5

0.43 NAc 5.14 1.53 NAd 1.10

0.38 NAc 0.99 3.72 NAd 0.63

NAe NAe NAe NAe NAe NAe

Free CN by microdiffusion (ASTM D4282-95)

DIW CW1 CW2 CW3 CW4 CW5

0.44 1.16 1.15 0.00g 0.53 0.87

0.85 0.75 0.00g 0.42 2.01 1.50

NAe NAe NAe NAe NAe NAe

Metal cyanides by ion chromatography

DIW CW1 CW2 CW3 CW4 CW5

NAb NAb NAb NAb NAb NAb

1.82 10.4 4.02 8.27 10.7 NAc

0.81 0.94 0.49 1.22 0.24 NAc

Total cyanide by low-power UV digestion

DIW CW1 CW2 CW3 CW4 CW5

0.42 NAf 0.73 1.34 0.43 NAc

0.99 NAf 0.37 0.88 0.36 NAc

0.86 NAf 0.22 0.69 0.36 NAc

Total cyanide by UV digestion and thin-film distillation (ASTM D4374-00)

DIW CW1 CW2 CW3 CW4 CW5

2.47 NAc 2.17 1.19 NAc NAc

1.68 NAc 1.53 2.47 NAc NAc

1.19 NAc 6.50 0.00 NAc NAc

Weak acid dissociable cyanide (SM 4500-CN I)

DIW CW1 CW2 CW3 CW4 CW5

0.72 NAc 0.87 0.53 0.88 1.68

0.64 NAc 1.05 0.64 1.17 1.76

NAe NAe NAe NAe NAe NAe

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TABLE 7.7 Continued Method detection limits (ppb) Method Weak acid dissociable cyanide by thin-film distillation (SM 4500-CN I, automated)

Water matrix

NaCN

K2 Ni(CN)4

K4 Fe(CN)6

DIW CW1 CW2 CW3 CW4 CW5

0.30 NAc 0.36 0.81 0.43 0.49

0.46 NAc 0.49 0.47 0.54 1.00

NAe NAe NAe NAe NAe NAe

a DIW = Reagent (deionized) water meeting the Type II specifications as defined in

ASTM D 1193-91 [69]. b NA = Not available because NACN is the main component in the eluent for ion chro-

matography technique. c NA = Not available because the measured level of analyte in the sample was greater

than five times the detection limit. d NA = Not available because untreated sulfide interferences prevented samples from being analyzed. e NA = Not available because this method is not intended to detect iron cyanide. f NA = Not available because the study on the MGP site groundwater was not conducted. g The MDL of 0.00 resulted from identical results in the seven replicates, yielding a standard deviation SD = 0.00, thus a MDL = 0.00. Source: Data from Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Environ. Sci. Technol., 37, 107, 2003.

significant interference problems and low recoveries. Additional sample pretreatment could have mitigated these problems, but since equivalent sample pretreatment was implemented for each method studied, the results reflect the robustness of the various techniques. There was significant recovery (76%) of diffusible cyanide in microdiffusion tests with nickel–cyanide-spiked samples, reflecting dissociation of this weak metal–cyanide complex and demonstrating that the test can recover some fraction of WAD cyanide in addition to free cyanide. Among the three WAD cyanide methods evaluated by Zheng et al. [23], the conventional method involving manual distillation, along with its automated version, the WAD thin-film distillation method, performed the best, with acceptable recoveries in all water matrices and excellent accuracy as determined from analysis of certified standards. The manual distillation WAD method can be performed with inexpensive instrumentation that is already available in many laboratories. The automated WAD method involving thin-film distillation can reduce significantly the time for analysis but requires specialized equipment. The available cyanide method provides for rapid analysis but requires specialized equipment and careful sample screening for potential interferences. Overall, however, the performance of all three methods was satisfactory and sufficient to recommend them as alternatives to the problematic CATC method for measurement of low concentrations of weakly complexed cyanide. Interest and use in the available cyanide technique is growing, primarily due to the speed and accuracy with which results can be obtained. The performance of the available cyanide method has been demonstrated successfully with other industrial wastewaters, including wastewater from a chemical production facility [27], and with gold mine tailings solutions [59,60]. In these studies, the potential for sulfide to cause significant interference was identified repeatedly, but it was also demonstrated that the sulfide interference can be addressed through appropriate pretreatment.

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TABLE 7.8 Quality Control Acceptance Criteria for USEPA Water Methods for Cyanide Measurement

USEPA Method 335.2 335.3 335.1 OIA-1677

Calibration verification recovery range (%) Not prescribed Not prescribed Not prescribed 86–118

Initial precision and recovery (IPR) Required recovery range (%) 45–125 64–136 64–136 92–122

Precision 40% RSD 36% RSD 36% RSD <5.1% RSD

Ongoing precision and recovery (OPR) Required recovery range (%) 40–130 60–140 60–140 82–132

Matrix spike/matrix spike duplicate

Precision

Required recovery range (%)

Precision

NAa NA NA NA

40–130 60–140 60–140 82–130

40% RPD 36% RPD 36% RPD <11% RPD

a NA = no data available.

Source: Data from USEPA, Method OIA-1677: Available cyanide by flow injection, ligand exchange and amperometry, EPA-821/R-99-013, U.S. Environmental Protection Agency, Office of Water, Washington, DC, 1999; and from USEPA, Guide to flexibility and approval of EPA water methods (draft), U.S. Environmental Protection Agency, Office of Water, Engineering and Analysis Division, Washington, DC, 1996.

7.10 QUALITY CONTROL CRITERIA FOR MEASUREMENT OF CYANIDE IN WATER All water analyses should be conducted in the framework of systematic quality assurance (QA) and quality control (QC) programs. A quality assurance program specifies the procedures required to produce defensible data of known precision and accuracy. Elements of a QA program include, for example, procedures for handling and receiving samples, standard operating procedures for each analytical method, and procedures for calibration of equipment, procedures for assessing data precision and accuracy [61]. A quality control program is relevant to a particular analytical method. Elements of a QC program include, for example, definition of QC indicators and QC acceptance criteria, calibration and recalibration, and analyses of reagent blanks, spiked reagent blanks, matrix spikes, and internal standards [61]. USEPA has promulgated standardized quality assurance/quality control acceptance criteria for cyanide measurement methods approved for use in water programs [62]. These methods include the following (i) Method 335.2, total cyanide analysis by distillation with spectrophotometric detection [63]; (ii) Method 335.3, automated total cyanide method [16]; (iii) Method 335.1, cyanide amenable to chlorination by distillation followed by spectrophotometric or titrimetric detection [64]; and (iv) Method OIA-1677, available cyanide by flow injection, ligand exchange, and amperometry [20]. Table 7.8 lists the critical quality control acceptance criteria for these methods, including the precision and recovery ranges. Guidance on quality control acceptance criteria is provided in descriptions for other cyanide analysis methods, such as those given in Standard Methods [3]. Performance goals and acceptance criteria need to be developed on a project-specific basis.

7.11 MEASURING CYANIDES — A REGULATORY DILEMMA A fundamental difficulty with use of the various analytical methods for cyanide in environmental science and engineering practice is that most of the analytical methods are not targeted on specific

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cyanide species, but on groups of compounds, while regulations are sometimes focused on particular species. For example, the U.S. drinking water maximum contaminant level (MCL) for cyanide is specified in terms of free cyanide [53], but the list of acceptable analytical methods for use in drinking water monitoring [53] includes the CATC method which does not report only free cyanide. A similar situation exists with respect to the U.S. ambient water quality criteria [43]. The freshwater and marine criteria are specified in terms of free cyanide, but the conventional cyanide-group methods such as total cyanide, WAD cyanide, and CATC are often used in water quality monitoring. This has led to much confusion in using different types of cyanide measurements in regulatory compliance contexts. The difficulty with measurement of cyanide in water and wastewaters for regulatory compliance purposes is a long-standing problem. In the United States, the problem goes back at least as far as when the list of toxic pollutants for water was being created in 1973. At that time, the USEPA’s Water Program proposed toxic pollutant effluent standards [65] as required by section 307(a)(1) of the Federal Water Pollution Control Act Amendments of 1972. The definition of cyanide at paragraph 129.05a, p. 35393, was given as: “Cyanide means any cyanide compound that will produce free cyanide ion or molecular hydrogen (HCN) in the effluent. The recommended analytical method should detect only simple cyanides after all interferences have been eliminated or their effects minimized by appropriate removal techniques.” A problem arose immediately in response to this notification because of an inconsistency in the proposed definition of cyanide as cited for simple cyanides and the recommended analytical procedures, which were for total cyanide. Industry submitted various affidavits in March, 1974 pointing out the inconsistency [66–68]. Industry wanted to be regulated on the basis of toxic forms of cyanide, not on all cyanide complexes and compounds. It appears that the subsequent change in the naming of the toxic pollutants from “cyanide and all cyanide compounds” to “cyanides” was an accommodation of that interest. Since then, there has been confusion over what species are being regulated and what is the appropriate analytical technique for regulatory compliance. The crux of this regulatory dilemma in the United States is the fact that modern treatment and toxicity studies are based on investigation of individual cyanide species, while many federal and state water quality regulations are based on what a specified analytical method reports rather than on the particular cyanide species to be regulated.

7.12 SUMMARY AND CONCLUSIONS • Cyanide is often released as an industrial by-product and is present as a dissolved contaminant in groundwater, surface water, and wastewater systems. • Most cyanide analysis methods for water are aimed at recovering inorganic forms of cyanide. Most existing analytical methods for inorganic cyanides provide for measurement of a group of species. These methods include total cyanide, weak acid dissociable cyanide (WAD), cyanide amenable to chlorination (CATC), available cyanide by ligand exchange, and free cyanide. Total cyanide recovers free cyanide (HCN, CN− ) and metal-complexed forms of inorganic cyanide; WAD, CATC, and available cyanide all measure weak metal– cyanide complexes plus free cyanide; free cyanide techniques usually measure the sum of HCN and CN− . • Total cyanide is the most common cyanide analysis measurement performed. • Of all three methods aimed at recovering weak metal–cyanide complexes, the CATC method is the least reliable method, as it has poor precision, bias, and recovery; the WAD and available cyanide methods are more reliable. • The analytical methods aimed at measuring individual species of cyanide are free cyanide by gas chromatography, gas–liquid diffusion (including microdiffusion), ion-selective electrode, and direct colorimetric analysis; metal–cyanide complexes by ion exchange chromatography and reversed-phase ion-pair chromatography; SCN− by colorimetric

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• • •

•

analysis; CNO− by hydrolysis and ion-selective electrode; and CNCl by colorimetric analysis. The techniques used, most commonly, are microdiffusion for free cyanide, and ion exchange chromatography for metal–cyanide complexes. For most cyanide analysis methods, sample screening for interferences and pretreatment to remove interferences is important. Common and especially important interferents of concern are oxidizing agents (e.g., chlorine) and sulfide. Available analytical methods yield only partial and unpredictable recovery of organocyanide compounds. Further method development is necessary for measurement of organocyanide compounds in water. From performance testing in waters spanning a broad range of composition, it has been determined that the existing cyanide analytical methods for inorganic cyanide are capable of low method detection limits (<10 ppb) with satisfactory accuracy and precision for their target analytes or analyte groups. The current use of the term “cyanide(s)” in regulations is hampered by not defining specific cyanide species being regulated. “Cyanide(s)” is often defined only by inference from the analytical procedure specified or chosen, and whatever the method reports.

REFERENCES 1. ASTM, Designation D 6696-01: Standard guide for understanding cyanide species, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2001. 2. Zheng, A., Dzombak, D.A., and Luthy, R.G., Effects of thiocyanate on the formation of free cyanide during chlorination and UV disinfection of POTW secondary effluent, Water Environ. Res., 76, 205, 2004. 3. APHA/AWWA/WEF, Method 4500-CN Cyanide, in Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S., Greenberg, A.E., Eaton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 4. Xie, Y. and Hwang, C.J., Cyanogen chloride and cyanogen bromide analysis in drinking water, in Encyclopedia of Analytical Chemistry, Meyers, R.A., Ed., John Wiley & Sons, Chichester, UK, 2000, p. 2333. 5. Yi, Y., Theis, T.L., and Young, T.C., The effect of chlorination on organocyanide compounds, Water Environ. Res., 74, 51, 2002. 6. Luthy, R.G. and Bruce, S.G., Kinetics of reaction of cyanide and reduced sulfur species in aqueous solution, Environ. Sci. Technol., 13, 1481, 1979. 7. Luthy, R.G., Bruce, S.G., Walters, R.W., and Nakles, D.V., Cyanide and thiocyanate in coal gasification wastewaters, J. Water Poll. Control Fed., 51, 2267, 1979. 8. Broderius, S.J. and Smith, L.L., Direct photolysis of hexacyanoferrate complexes: proposed applications to the aquatic environment, EPA-600/3-80-003, U.S. Environmental Protection Agency, Duluth, MN, 1980. 9. Johnson, C.A., Leinz, R.W., Grimes, D.J., and Rye, R.O., Photochemical changes in cyanide speciation in drainage from a precious metal ore heap, Environ. Sci. Technol., 36, 840, 2002. 10. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 11. Scott Rader, W., Solujic, L., Milosavljevic, E.B., and Hendrix, J.L., Sunlight-induced photochemistry of aqueous solutions of hexacyanoferrate-(II) and -(III) ions, Environ. Sci. Technol., 27, 1875, 1993. 12. ASTM, Designation D 2036-98: Standard test methods for cyanides in water: method a total cyanides after distillation, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 13. USGS, Method I-4302-85: Cyanide, total recoverable; colorimetric, barbituric acid automatedsegmented flow, in USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, 1985, available at http:www/nemi.gov.

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14. USGS, Method I-3300-85: Cyanide, total, colorimetric, pyridine–pyrazolone, in USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, 1985, available at http://www.nemi.gov. 15. USEPA, Method 335.4: Determination of total cyanide by semi-automated colorimetry, Rev. 1993, in Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, 1979, available at http://www.nemi.gov. 16. USEPA, Method 335.3: Cyanide, total (colorimetric, automated UV), in Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, 1979, available at http://www.nemi.gov. 17. USEPA, Technical notes on drinking water methods, EPA-600/-R-94-173, U.S. Environmental Protection Agency, Washington, DC, 1994. 18. ASTM, Designation D 4374-00: Standard test methods for cyanides in water: automated methods for total cyanide, dissociable cyanide and thiocyanate, in Annual Book of ASTM Stantards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2000. 19. Kelada, N., Automated direct measurements of total cyanide species and thiocyanate, and their distribution in wastewater and sludge, J. Water Pollut. Control Fed., 61, 350, 1989. 20. USEPA, Method OIA-1677: Available cyanide by flow injection, ligand exchange and amperometry, EPA-821/R-99-013, U.S. Environmental Protection Agency, Office of Water, Washington, DC, 1999. 21. USEPA, Method OIA-1677: Available cyanide by flow injection with ligand exchange, Fed. Regist., 64, 73414, 1999. 22. Altmayer, F., Cyanide: ATC headaches, Plating Surf. Finish, 2, 26, 1997. 23. Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Evaluation and testing of analytical methods for cyanide species in municipal and industrial contaminated waters, Environ. Sci. Technol., 37, 107, 2003. 24. Ghosh, R.S., Muraka, I., Nakles, D.V., and Neuhauser, E.F., Refinement of weak acid dissociable (WAD) method for measuring weak metal–cyanide complexes in aqueous samples, Environ. Eng. Sci., 25, 543, 2005. 25. ASTM, Designation D 6888-03: Standard test method for available cyanide with ligand displacement and flow injection analysis (FIA) utilizing gas diffusion separation and amperometric detection, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2003. 26. Milosavljevic, E.B., Solujic, L., and Hendrix, J.L., Rapid distillationless “free cyanide” determination by a flow injection ligand exchange method, Environ. Sci. Technol., 26, 511, 1995. 27. Sebroski, J.R. and Ode, R.H., Method comparison and evaluation for the analysis of weak acid dissociable cyanide, Environ. Sci. Technol., 31, 52, 1997. 28. Carr, S.A., Baird, R.B., and Lin, B.T., Wastewater derived interferences in cyanide analysis, Water Res., 31, 1543, 1997. 29. Zheng, A., Dzombak, D.A., and Luthy, R.G., Formation of free cyanide and cyanogen chloride from chlorination of POTW secondary effluent: laboratory study with model compounds, Water Environ. Res., 76, 113, 2004. 30. Otu, E.O., Byerley, J.J., and Robinson, C.W., Ion chromatography of cyanide and metal cyanide complexes: a review, Intern. J. Environ. Anal. Chem., 63, 81, 1996. 31. USEPA, Method 9015: Metal cyanide complexes by anion exchange chromatography and UV detection, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, U.S. Environmental Protection Agency. Office of Solid Waste, Washington, DC, 2004, Rev 0. 32. ASTM, Designation D 6994-04. Standard test method for determination of metal cyanide complexes in wastewater, surface water groundwater and drinking water using anion exchange chromatography with UV detection, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2004. 33. Dionex, Determination of Metal Cyanides, Application Note 55, Dionex Corporation, Sunnyvale, CA, 1988. 34. Grigorova, B., Wright, S.A., and Josephson, M., Separation and determination of stable metallo-cyanide complexes in metallurgical plant solutions and effluents by reversed-phase ion-pair chromatography, J. Chromatogr., 410, 419, 1987.

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35. Haddad, P.R. and Rochester, N.E., Ion-interaction reversed-phased chromatographic method for the determination of gold (I) cyanide in mine process liquors using automated sample preconcentration, J. Chromatogr., 439, 23, 1988. 36. Hilton, D.F. and Haddad, P.R., The determination of metal cyano complexes by reversed-phase ioninteraction high performance liquid chromatography and its application to the analysis of precious metals in gold processing solutions, J. Chromatog., 361, 141, 1986. 37. Huang, Q., Paull, B., and Haddad, P.R., Optimisation of selectivity in the separation of metallo-cyanide complexes by ion-interaction liquid chromatography, J. Chromatogr., 770, 3, 1997. 38. Waters, Metal Cyano Complex Analysis, Application Method 304, Waters Corporation, Milford, MA, 1990. 39. Haddad, P.R. and Kalambaheti, C., Advances in chromatography: speciation of microgram per liter levels of metallo-cyandes using ion-interaction chromatography, Anal. Chim. Acta, 250, 21, 1991. 40. Claeys, R.R. and Freund, H., Gas chromatographic separation of HCN on Poropak Q, Environ. Sci. Technol., 2, 458, 1968. 41. Schneider, C.R. and Freund, H., Determination of low level hydrocyanic acid in solution using gas–liquid chromatography, Anal. Chem., 34, 69, 1962. 42. Doudoroff, P., Leduc, G., and Schneider, C.R., Acute toxicity to fish of solutions containing complex metal cyanides, in relation to concentrations of molecular hydrocyanic acid, Trans. Am. Fish. Soc., 95, 6, 1966. 43. USEPA, Ambient water quality criteria for cyanide-1984, EPA-440/5-84-028, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 1984. 44. APHA/AWWA/WEF, Method 207: Cyanide, in Standard Methods for the Examination of Water and Wastewater, 13th ed., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1971. 45. Dzombak, D.A. and Higgins, C.J., Comparison of direct colorimetric analysis with microdiffusion for measurement of free cyanide, Quarterly Progress Report (May 27), WERF Project 01-ECO-1, Water Environment Research Federation, Alexandria, VA, 2004. 46. ASTM, Designation D 4282-95. Standard test method for determination of free cyanide in water and wastewater by microdiffusion, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 47. Kruse, J.M. and Thibault, L.E., Determination of free cyanide in ferro- and ferricyanides, Anal. Chem., 45, 2260, 1973. 48. ANSI, Method for determining free cyanide in photographic effluents, in ANSI PH 4.41, American National Standards Institute, New York, 1978. 49. Broderius, S.J., Determination of hydrocyanic acid and free cyanide in aqueous solution, Anal. Chem., 53, 1472, 1981. 50. Broderius, S.J., Determination of molecular hydrocyanic acid in water and studies of the chemistry and toxicity to fish of metal–cyanide complexes. Ph.D. thesis, Department of Fisheries and Wildlife, Oregon State University, Corvallis, OR, 1973. 51. NGA, Investigation of analytical methods, aqueous speciation and toxciology of cyanide at manufactured gas plant (MPG) sites in State of New York, Northeast Gas Association, New York, 2004. 52. USEPA, SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. 53. USEPA, National primary drinking water regulations. Section 141.23: Inorganic chemical sampling and analytical requirements, in Code of Federal Regulations, Title 40, U.S. Environmental Protection Agency, Washington, DC, 2002. 54. Sekerka, I. and Lechner, J.F., Potentiometric determination of low levels of simple and total cyanides, Water Res., 104, 479, 1976. 55. Xie, Y. and Reckhow, D.A., A rapid and simple analytical method for cyanogen chloride and cyanogen bromide in drinking water, Water Res., 27, 507, 1993. 56. Seigler, D.S., Isolation and characterization of natually occurring cyanogenic compounds, Phytochemistry, 14, 9, 1974. 57. Theis, T.L., Young, T.C., and Yi, Y., Evaluation, development and testing of analytical methods, Quarterly Progress Report (Nov. 30), WERF Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 1999.

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58. Schaefer, R.J., Photocatalytic treatment of cyanide in aluminum potlining leachate using ozone as an oxidizing agent. M.S. thesis, Clarkson University, Potsdam, NY, 1996. 59. Evans, J.D., Thompson, L., Clark, P.J., and Beckman, S.W., Method comparison study for weak acid dissociable cyanide analysis, Environ. Sci. Technol., 37, 592, 2003. 60. Milosavljevic, E. and Solujic, L., How to analyze for cyanide, in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G. and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 117. 61. APHA/AWWA/WEF, Method 1020: Quality assurance and quality control, in Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S., Greenberg, A.E., and Eaton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 62. USEPA, Part 136 — Guidelines establishing test procedures for the analysis of pollutants, in Code of Federal Regulations, Title 40, Sec. 136.3, U.S. Environmental Protection Agency, Washington, DC, 2004. 63. USEPA, Method 335.2: Cyanide, total (titrimetric, spectrophotometric), Rev. 1980, in Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, 1979, available at http://www.nemi.gov. 64. USEPA, Method 335.1: Cyanides amenable to chlorination (titrimetric, spectrophotometric), in Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, 1979, available at http://www.nemi.gov. 65. USEPA, Proposed toxic pollution effluent standards, 40 CFR Part 129, Fed. Regist., 38, 35388, 1973. 66. Caruso, S.C., Comments on proposed toxic pollutant effluent standards, representing American Iron and Steel Institute, Affidavit submitted to USEPA, FWPCA (307), Docket No. 1, U.S. Environmental Protection Agency, Washington, DC, 1974. 67. Doudoroff, P., Comments on proposed toxic pollutant effluent standards, representing American Iron and Steel Institute, Affidavit submitted to USEPA, FWPCA (307), Docket No. 1, U.S. Environmental Protection Agency, Washington, DC, 1974. 68. Young, E.F., Comments on proposed toxic pollutant effluent standards, representing American Iron and Steel Institute, Affidavit submitted to USEPA, FWPCA (307), Docket No. 1, U.S. Environmental Protection Agency, Washington, DC, 1974. 69. ASTM, Designation D 1193–91. Standard Specification for reagent water, Annual Book of ASTM Standards, Vol. 11.01, ASTM International, West Conshohocken, PA, 1998.

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of Cyanide in Solids 8 Analysis and Semi-Solids David A. Dzombak, Joseph T. Bushey, Sharon M. Drop, and Rajat S. Ghosh CONTENTS 8.1 8.2 8.3 8.4 8.5 8.6 8.7 8.8

Direct Acid Distillation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acid Solution Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Neutral Solution Extraction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Alkaline Solution Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Extraction Solution pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Liquid–Solid Ratio . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effect of Extraction Time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Extraction of Plant and Animal Tissue, and other Semi-Solid Materials . . . . . . . . . . . . . . . . . 8.8.1 Plant Tissue Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.8.2 Animal Tissue Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.8.3 Alternative Extraction Procedures for Plant Tissue . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.8.4 Analytical Methods for Use with Plant Tissue . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.9 Quality Control and Method Performance Evaluation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.10 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

156 157 158 158 159 161 162 163 163 164 164 164 165 165 166

Cyanide can be associated with solids in contaminated water and soil systems. Therefore, methods for analyzing total cyanide or specific cyanide species in soils, sludges, plant and animal tissue, solid wastes, and other solids are of interest. As discussed in Chapters 2 and 5, free cyanide is very reactive and readily forms solution complexes and solid phase precipitates with metals in the environment. It can form solid precipitates with a number of metals, including copper, zinc, nickel, and iron among others [1]. Iron–iron cyanide solids such as Prussian Blue, Fe4 (Fe(CN)6 )3 (s), and Turnbull’s Blue, Fe3 (Fe(CN)6 )2 (s), are the most common cyanide solid phases in the environment, as iron is very reactive with free cyanide, and is ubiquitous [2]. Iron–iron cyanide solids have limited to moderate solubility in water, with the solubility dependent on pH, and oxidation–reduction potential [2,3]. Solid phases consisting of simple salts of free cyanide or metal–cyanide complexes, such as KCN, NaCN, and K2 Ni(CN)4 , are not encountered in the environment to any appreciable extent, as these solids dissolve readily in water. Aqueous cyanide species can also adsorb onto various inorganic and organic phases, with the extent of adsorption being dependent on pH, and other solution conditions [4–9]. Thus, adsorption is another means by which cyanide species can become associated with the solid phase. In addition to precipitated and adsorbed cyanide species in soils and sludges, cyanide can also be present in entrapped or bound water in drained samples of these materials. The existence of cyanide 155

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in the aqueous phase soils and sludges needs to be considered when analyzing for cyanide in samples of these “solids.” Analysis of cyanide content of solids involves aqueous dissolution of the cyanide species of interest using acidic, neutral, or caustic solution, followed by separation of the residual solids, and then by analysis of the aqueous extracts. Any of the aqueous phase analyses discussed in Chapter 7 can be applied to the aqueous phase extract. As will be shown, selection of a low or mid-range pH for the extraction can result in incomplete recovery of cyanide, depending on the form of the cyanide present in the solid, and the manner in which it is associated with the solid. Selection of the extraction condition depends on the goals of the analysis, for example, maximum recovery of cyanide from the sample, or leachability of cyanide from the sample under particular conditions. A key issue in analysis of solids for cyanide or any other contaminant, is that of acquisition of representative samples for analysis. Heterogeneity in soil at a contaminated site, in sludge deposited in a tank or pond, among identical species of plants in a field, or solid wastes in a fill, is an issue that has to be addressed in the development of any sampling program. Obtaining representative samples is a significant challenge that often demands rigorous statistical analysis. Because this issue is not unique to analysis of cyanide, it is not examined further, but in thinking about solids analysis, one is well advised to always keep in mind the importance of the sampling plan [10–13]. In this chapter, several standard techniques for extraction of cyanide in solids are reviewed, and types of samples for which each is most appropriate are discussed. Results from the application of these methods to different types of cyanide-bearing solids are examined and serve as the basis of recommendations for extraction and analysis approaches to use when analyzing cyanide in solids.

8.1 DIRECT ACID DISTILLATION The common Standard Methods [14] technique for total cyanide analysis in water, which involves distillation of HCN from acidified solution, has provisions for analysis of “insoluble cyanide” in solids that entails acid leaching. For analysis of “insoluble cyanide,” one approach specified in Standard Methods is to place a 500 mg sample of the solids into a distilling flask. This is followed by addition of 500 ml of deionized water, and then acidification (sulfuric acid) and distillation in the usual manner for analysis of total cyanide in water. That is to say, this approach involves placing a sample of the cyanide-bearing solids directly in the distillation flask and liberating all cyanide that can be extracted under strongly acidic conditions in the presence of heat. Direct acid distillation for analysis of total cyanide in solids has also been specified by the U.S. Geological Survey (USGS) [15,16]. USGS Method I-6302-85 [15] employs manual sample distillation and pyridine-barbituric acid for colorimetric analysis as used in Standard Methods [14]. USGS Method I-5300-85 [16] employs pyridine–pyrazolone for colorimetric analysis, which was specified in older versions of Standard Methods (e.g., the 13th edition [17]). The USEPA [18] has specified in its SW-846 manual a similar direct-acid distillation procedure for cases in which the total cyanide concentration in the solids is sufficiently high (defined as greater than about 50 µg/g) so that detectable cyanide is yielded. In this method, it is specified that 1 to 10 g of solid be placed in 500 ml of deionized water in the distillation flask, followed by acidification, and distillation in the usual manner for analysis of total cyanide in water [19]. Mansfeldt and Biernath [20] developed a modification to the standard German distillation technique for cyanide (similar to the Standard Methods total cyanide procedure) that involves placement of 100 to 500 mg of soil in 10 ml of 0.1 M HCl for 2 h at 108◦ C. The free cyanide released from the heated, acidified soil suspension is trapped in NaOH solution, which is analyzed for total cyanide content. A side-by-side comparison of the performance of the microdistillation procedure with the conventional, larger volume distillation procedure was conducted using soil samples from coke plants and blast furnace sludge samples. It was demonstrated that the microdistillation procedure produced results equivalent to those obtained with the larger volume distillation procedure.

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Since acidic conditions generally favor adsorption of metal–cyanide complexes on soils and other materials bearing oxidic solids, and are not favorable for dissolution of some metal–cyanide solids (e.g., iron–iron cyanide solids [2]), direct acid distillation and analysis of cyanide solids will in, many cases, result in incomplete recovery of total cyanide when metal–cyanide complexes are present. Moreover, such analysis often leads to irreproducible recoveries when replicate samples are tested. In such cases, aqueous extraction (leaching followed by filtration of residual solids) with a neutral or alkaline solution is a critical step that must precede acid distillation to ensure quantitative recovery of metal–cyanide complexes such as iron-cyanide species. Depending on the cyanide species to be determined, extraction solutions of either neutral or alkaline pH concentrations may be used. Extraction methods are discussed below.

8.2 ACID SOLUTION EXTRACTION It is also possible to leach solids with acid under room temperature conditions, and then subject the filtered extract to aqueous phase analysis for total cyanide or other forms of cyanide. This is done, for example, in cases where information about leaching under acidic environmental conditions is desired. Precautions need to be taken, such as zero headspace leaching in a closed vessel, to prevent HCN loss, as much as possible, during the extraction. Various kinds of acid leach tests have been performed. Theis et al. [21] conducted leach tests on manufactured gas plant (MGP) cyanide-bearing wastes for pH values ranging from 3 to 11. Acidic conditions prevail in waste disposal areas at many MGP sites [21,22]. Nitric acid (HNO3 ) solutions were used for the low pH leach tests. MGP waste samples and acid extract solution were placed in closed, pH-controlled, stirred reactors, and allowed to react for different periods of time. At the end of the designated time, a portion of the suspension was withdrawn and filtered, and the extract was analyzed for total cyanide, and a number of other chemical constituents. Results for an MGP waste sample (48 h reaction period) are presented in Figure 8.1. As seen there, some cyanide was extracted from the sample in the range pH 3 to 5, with the amount extracted dependent on pH. Contact of a soil, sludge, or other solid material with strong acid will usually result in the dissolution of many chemical constituents, as is apparent in Figure 8.1. This can lead to extract solutions of very complex composition, making some cyanide analytical methods difficult to apply to the extract solution. Methods that are more sensitive to water matrix interferences, such as ion selective electrode and ion exchange chromatography, will not be useful for direct analysis of many extract solutions. On the other hand, the methods involving initial distillation of the extract solution, such as the total cyanide method, work well even with complex matrices. The USEPA Toxicity Characteristic Leaching Procedure [23], and the similar ASTM test [24], employs a weak (acetic) acid extract solution. Cyanide is not one of the leachate constituents included in the TCLP methods development and is not specified as a TCLP extract solution analyte. Nevertheless, TCLP extract solutions have sometimes been analyzed for total cyanide or other forms of cyanide. This is inappropriate, as free cyanide is lost by volatilization in the conventional TCLP test. It may be possible to employ the zero headspace version of the TCLP test with higher recovery of cyanide, but this has not been demonstrated through published method development. Acid leaching can be effective in extracting free cyanide from soils, sludges, and wastes, but as is evident in Figure 8.1, acid leaching can result in low extraction yields for metal-complexed cyanides existing within solids in adsorbed or precipitated form. For example, dissolved iron cyanide adsorbs most strongly onto oxidic minerals in soils at pH < 6 [4], and iron-cyanide solids exhibit lowest solubility at low pH values [2,25]. Thus, if the form of the cyanide in the solids is unknown, solids containing cyanides should be subjected to neutral and alkaline solutions in addition to acid solution to assess maximum recoverable cyanide.

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104

104

NH3 Fe

Concentration, mg/g dry wt.

3

103

CN

10

102

102 Cu Ni

10

10

Co

1

1 Pb

10–1

10–1

Zn 2

4

6

8

10

12

pH

FIGURE 8.1 Results of acid, neutral, and alkaline extraction tests on an MGP site soil for total cyanide, ammonia, and selected metals. Reaction period = 48 h. (Source: Reprinted from: Theis, T.L. et al., Environ. Sci. Technol., 28, 99, 1994. Copyright 1994, American Chemical Society.)

8.3 NEUTRAL SOLUTION EXTRACTION To analyze for readily soluble cyanide in solids, such as solids with entrapped water containing free cyanide, or containing crystals of free cyanide salts such as KCN or NaCN, neutral-pH, deionized water is often used. Standard Methods [14] specifies placement of 500 mg of solids in 500 ml of distilled water and one hour of stirring for extraction of “soluble cyanide,” with aqueous phase cyanide analysis then performed on the extract solution. As noted in Standard Methods, low cyanide concentration in the extract solution may indicate the presence of “insoluble metal cyanides.” Extraction with neutral pH solution may also be performed to assess cyanide that is extractable from solids under rainwater contact conditions, such as at contaminated soil sites. The USEPA Synthetic Precipitation Leaching Procedure [26], and the similar ASTM method [27], was designed for such an assessment, and employs reagent (deionized) water. Cyanide is not one of the extract solution analytes routinely analyzed in the SPLP test, but the method description indicates that the test can be used to measure water extractable cyanide (“Extraction Fluid 3 — reagent water” is recommended for cyanide). There does not appear to be any specific method development for cyanide, so the potential for loss of free cyanide through volatilization is likely, similar to that in the TCLP test. It may be possible to employ the zero headspace version of the SPLP test with higher recovery of cyanide, but this does not appear to have been demonstrated through method development.

8.4 ALKALINE SOLUTION EXTRACTION In the Standard Methods [14] technique for total cyanide analysis in water, an alternative to the direct acid distillation approach for analysis of “insoluble cyanide” in solids, involves leaching for 12 to 16 h with 10% (2.5 M) NaOH solution. A 500 mg sample of the cyanide-bearing solids is

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placed in 500 ml of the extract solution and the suspension is mixed for 12 to 16 h. At the end of that time, the extract solution is separated from the solids, and then subjected to aqueous phase cyanide analysis. No specifications are provided in Standard Methods for the type of mixing, light control (exclusion), or the method of solid–liquid separation. Gravity settling is usually sufficient for solid– liquid separation. Filtration (0.45 µm) can be employed if the extract solution contains a significant colloidal fraction, but may increase cyanide loss from solution by volatilization of free cyanide [28], though volatilization losses should be small for high pH solutions in which the cyanide ion dominates speciation. Exclusion of light is a significant issue for samples containing metal–cyanide complexes subject to photodissociation, such as dissolved iron–cyanide complexes [3,25]. The USEPA method for analysis of cyanide in solids [18] also provides for an alkaline solution extraction prior to analysis as an alternative to direct acid distillation. As discussed in the method description, the alkaline (NaOH) extraction is intended for use when cyanide concentrations or forms in the solids are such that large amounts of soil would be required to yield detectable cyanide via direct-distillation acid leach. The method specifies that 25 g of solids be placed in 500 ml of reagent water adjusted to pH 12 with 5 ml of 50% w/w NaOH (giving an extract solution NaOH concentration of 0.25 M), and extracted for 16 h. To maintain pH > 10 during the extraction, 5 ml aliquots of 50% w/w NaOH are added as needed. Variations of the alkaline extraction techniques described above have been employed by a number of investigators. Mansfeldt and coworkers employed an alkaline extraction method to leach cyanide from blast furnace sludge [29] and from contaminated soil near a former coking plant [30]. The extraction procedure involves dispersing 10 g of dry solids in 250 ml of 1 M NaOH, closing the reaction vessel and placing it on an end-over-end rotator, allowing reaction for 16 h, and then centrifuging for solid–water separation. Extract solution is then analyzed for total cyanide using a microdistillation procedure [20]. In an alkaline extraction procedure developed by Meeussen et al. [25], 5 g of contaminated soil is dispersed in 100 ml of 0.25 M NaOH and the suspension is heated on a boiling water bath for three hours. Subsequently, the suspension is filtered and the aqueous phase is analyzed for total and free cyanide.

8.5 EFFECT OF EXTRACTION SOLUTION pH Extraction of cyanide from solids with acid, base, or neutral pH aqueous solutions yields substantially different results for particular solids. The transfer of cyanide mass from the solid phase to the aqueous phase under any of these conditions depends on the form of cyanide in the solid phase. For example, if cyanide is present as a metal–cyanide complex adsorbed on oxide mineral components of a soil, higher pH conditions (e.g., pH > 7) are required to induce desorption of the metal cyanide species [4]. Extraction of the soil at low pH would yield very little or no dissolved metal–cyanide species. Similarly, if cyanide is present as a metal–cyanide solid, the solubility of the solid will be strongly dependent on pH [1,2]. Thus, the selection of an acid, base, or neutral solution extraction of solids will significantly affect the results obtained for solid phase cyanide analyses. These are not interchangeable extractants. Ghosh et al. [31], Köster [32], Theis et al. [21], and Mansfeldt et al. [33] examined the effect of leaching contaminated, manufactured gas plant (MGP) site solids over a range of pH values, as adjusted with NaOH or H2 SO4 addition. Ghosh et al. investigated extraction of cyanide from nine different MGP site contaminated soils under strongly acidic (pH 1.8), mildly acidic (pH 5.1), and strongly alkaline (pH > 12) conditions. All extractions were performed for 16 h. Köster investigated 8-day extractions with extract solutions having pH values 3 to 9, while Theis et al. examined pH 3 to 11 with 48 h of mixing and Mansfeldt et al. pH 3 to 12 with a 3-day extraction period. Each found that the measured cyanide content increased with pH, and used the result obtained with the highest pH solution for the total cyanide content of the solids. Meeussen et al. [34] also studied the cyanide content of MGP site solids, and used an extract obtained from soil in heated 0.25 M NaOH solution

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100,000 Total cyanide concentration, mg/Kg

Caustic leach (pH=14) DI Water leach (pH=5.1) 10,000

Acid leach (pH=1.8)

1000

100

10

1 A

B

C

D

E

F

H

I

MGP site ID

FIGURE 8.2 Total cyanide concentration in nine MGP site soils as determined from leach tests with acid (pH 1.8), deionized water (pH 5.1), and NaOH (pH 14). (Source: Ghosh, R.S. et al., Environ. Eng. Sci. 21, 752, 2004. With permission.)

for 3 h to determine total cyanide. The MGP soil leach test results of Theis et al. [21], shown in Figure 8.1, illustrate well the increase in amount of recoverable cyanide with increasing pH. Similar results were obtained by Ghosh et al. [31] for nine MGP site soils, as shown in Figure 8.2. In MGP site soils, the predominant form of cyanide is iron–iron cyanide solids, which under oxidizing conditions have maximum solubility at high pH [2]. Young and Theis [35] applied both the neutral-pH, reagent water extraction and the NaOH extraction specified in Standard Methods to soils from two MGP sites. The reagent water extractions yielded no detectable cyanide. In addition, free cyanide (NaCN) spiked into the samples was not recovered using the reagent water. As the MGP site soils were acidic and lowered the pH of the extraction solution, it was hypothesized that volatilization loss of HCN may have occurred during the extractions. In contrast, relatively high concentrations of cyanide were measured using the NaOH extraction solution. Further, the recovery of soil spike additions of NaCN and K4 Fe(CN)6 into the soil samples was high, with an average recovery of 80.6%. Bushey and Dzombak (unpublished data) compared the direct acid distillation and the 10% NaOH extraction method for an MGP site soil. The 10% NaOH solution was sonicated for 20 min prior to the 16-h leach to enhance contact with the soil. For the direct acid distillation method, the effect of performing repeated 2-h distillations, in an attempt to increase cyanide recovery, was also examined. The results are presented in Figure 8.3, where it may be seen that recovery of cyanide with the 10% NaOH extraction was an order of magnitude greater than that of direct acid distillation approach, even with repeated distillations. Mansfeldt and Biernath [29] analyzed cyanide content of blast furnace sludge samples by direct acid distillation (2-h distillation period) and by alkaline extraction with 1 M NaOH and a 16-h reaction period. The alkaline extraction was repeated three times in succession, to evaluate incremental recovery. The primary form of cyanide in the sludge was the crystalline cyanide-containing compound potassium zinc hexacyanoferrate (II) nonahydrate, K2 Zn3 [Fe(CN)6 ]2 ·9H2 O. From application of the two analytical methods to 32 blast furnace sludge samples, it was found that the two methods yielded comparable results. The potassium–zinc–iron cyanide salt was completely dissolved in both acid and base solutions. In the repeated alkaline extractions, it was found that 85% of the total recoverable cyanide was removed in the first extract, 12% in the second extract, and 3% in the third extract.

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Total cyanide concentration (mg as CN/kg)

6000

5000

4000

3000

2000

1000

0 NaOH Leach Distilled 3×

Distilled 4×

Distilled 5×

FIGURE 8.3 Comparison of total cyanide concentrations measured for an MGP site soil base extract (10% NaOH extraction for 16 h) and for the same soil subjected to direct-distillation acid leach with repeated 2-h distillations.

If alkaline extraction is employed, for application of certain analytical procedures on the extract solution, it is important to adjust the pH of the extract solution prior to analysis, so that it is not too high. A Clarkson University study (unpublished data) of the effect of alkaline extract solution [18] pH on the analysis of free cyanide in MGP site soils by microdiffusion [36] revealed that the pH of the leachate needs to be adjusted to 11 to 12 by dropwise addition of H2 SO4 . Similarly, Ghosh et al. [37] found it important to filter alkaline extract solutions [18] from MGP site and aluminum smelter waste solids and adjust the pH to 11 to 12 prior to subjecting the leachate to ion chromatographic analysis [38]. This pH adjustment enables proper extract sample conditions for application of the analytical methods cited. Considering the diversity of aqueous and solid phase cyanide species, and of the solid phase matrices in which cyanide may be adsorbed, precipitated, or otherwise retained, it is advisable to conduct acid, neutral, and alkaline solution extractions on a solid with unknown cyanide content. Analysis of these extract solutions for total cyanide or a particular cyanide species or group of species will provide comprehensive insight into the aqueous leachability of the cyanide-bearing solid. Further, the extract solutions can be analyzed for other chemical species to obtain information about species that may complex with cyanide or influence cyanide chemistry indirectly. In their study of leaching properties of MGP site soils, Theis et al. [21] analyzed the variable-pH extract solutions for trace metals, as well as, major cations and anions in addition to cyanide species (Figure 8.1). Ghosh et al. [31] studied the speciation of MGP site soil leachates under various pH conditions and found the ferrocyanide (Fe(CN)4− 6 ) species to dominate speciation under alkaline extraction conditions.

8.6 EFFECT OF LIQUID–SOLID RATIO The liquid–solid (LS) ratio employed can influence the efficiency of chemical extraction from solids in aqueous suspension [39]. The degree of the influence depends on the solid particle characteristics, and the form of the chemical. In cases where the concentration of the chemical is determined by dissolution to equilibrium, for example, the leachate concentration would not depend on LS ratio. However, for cases in which desorption or a mass-transfer-limited solid to water exchange process are involved, the leachate concentration does depend on LS ratio. Thus, selection of the LS ratio for

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Total cyanide concentration (mg as CN/kg)

2500

50 mg solid 200 mg solid 500 mg solid

2000

1500

1000

500

0 1

2

3

4

Soil

FIGURE 8.4 Effect of LS ratio on direct-acid distillation analysis of total cyanide in four MGP site soil samples. Each soil sample was distilled directly in 50 ml of strong acid solution. Three different solid amounts were used for each soil: 50, 200, and 500 mg. LS ratios (w/w) were thus 1000, 250, and 100.

cyanide extractions should be done with care. In the preceding sections, it was seen that the standard methods specify different LS ratios, and that some methods specify a range of acceptable LS ratios (e.g., that in Standard Methods [14]). Data for four MGP site soil samples obtained by Bushey and Dzombak (unpublished data) illustrate the effect of LS ratio in acid extraction of cyanide. Total cyanide analysis of the soils was performed directly using the Midi-Distillation procedure [40] which is a reduced-scale version of the Standard Methods [14] total cyanide distillation method. For direct acid distillation of solids, Standard Methods specifies an LS ratio of 1000:1 on a mass basis. To examine the effect of LS ratio, direct acid distillation was carried out using 50, 200, and 500 mg of solids (LS ratios of 1000, 250, and 100, respectively). As shown in Figure 8.4, it was found that acid extractable cyanide increased in proportion to LS ratio. While recoveries of acid extractable cyanide are relatively low from the iron– iron cyanide solids present in MGP site soils (see Figure 8.3), the recoveries were sufficient in these experiments to demonstrate the effect of LS ratio. Contaminated soils, sludges, and waste materials are heterogeneous and sufficient solids must be utilized in extractions to obtain representative analyses. However, too much solid relative to the extract volume can inhibit the extraction, which is evident in Figure 8.4. While the results shown in Figure 8.4 were obtained with acid extractions and results with neutral or alkaline extractions would be different, the data demonstrate that selection of LS ratio must be done with care. Also, solid sample homogenization can decrease variability in results obtained. For a sample with unknown properties, it would be useful to perform extractions at several LS ratios to obtain information about how leachable cyanide varies with LS ratio, or not.

8.7 EFFECT OF EXTRACTION TIME The extraction time required for quantitative analysis of cyanide will depend on the cyanide analyte of interest and the sample matrix. Standard Methods recommends an extraction period of 12 to 16 h for determining “insoluble cyanide” [14]. This is similar to the TCLP leaching procedure [23], which specifies an extraction time of 16 to 20 h. Lupichuk et al. [41] examined the effect of extraction time on total and free cyanide determination in aluminum reduction plant wastes. Samples were leached

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from 2 to 48 h under alkaline conditions at an LS ratio of 100. The leachates were filtered and analyzed for total cyanide [14] and free cyanide [36] content. The total and free cyanide concentrations in the extracts were found to be essentially equivalent for all leaching periods examined. Theis et al. [21] evaluated the effect of extraction time on MGP soil samples. Samples were extracted in solutions ranging from pH 7 to 11 at an LS ratio of 100 for varying time periods ranging from 0 to 187 h. The extracts were filtered and analyzed for total cyanide. In all cases, the total cyanide content was found to increase with extraction time. Overall, these studies indicate that the extraction time should be selected based on the cyanide species of interest. For free cyanide determination, shorter extraction times may be feasible. In the case of metal–cyanide complex salts, a minimum of 16-h extraction time is recommended. When determining total cyanide, more than one extraction time may be needed to confirm that all cyanide species present have been completely extracted from the sample.

8.8 EXTRACTION OF PLANT AND ANIMAL TISSUE, AND OTHER SEMI-SOLID MATERIALS Cyanide analysis in solids such as plant tissue, animal tissue, and sludges is complicated by high water content and by high organic content. The abundance of water presents the possibility of cyanide concentrating in the water phase rather than with the associated solid. Such samples differ from typical soil samples, which have cyanide associated with the solid surface through precipitation or adsorption. Drying of a high-water content sample prior to analysis is not an effective pretreatment, as this procedure increases the potential for volatilization of free cyanide. Recovery from wet samples is lowered upon drying, particularly if the cyanide is present in the free form [28]. Samples with high water content can be analyzed without drying via normal solid analysis techniques. However, care must be taken with units, with results typically presented on a “wet” basis. Separate analysis of the water content provides a means of adjusting the results to a “dry” basis. For high associated aqueous concentrations, the liquid cyanide fraction can also be accounted for, to obtain a measure of the cyanide associated with the solid. For plant and animal tissue, preventing sample contamination during analysis is particularly important due to the low concentrations typically associated with tissue samples. Generation of cyanide in acid digestion can occur via reactions involving organic material. While the cyanide mass generated may be small, concentrations of interest are usually low, often near the detection limit for analytical procedures such as total cyanide by distillation [14] and free cyanide by microdiffusion [36]. The release of cyanide from cyanogenic glycosides, as discussed in Chapter 3, or as a result of tissue destruction can artificially elevate cyanide results. Another concern is the potential release of organics or sulfides, both analytical interferences (see Chapter 7), into the extraction media. Yet, tissue destruction is necessary to bring about complete release to solution and subsequent analysis of tissue cyanide content.

8.8.1 PLANT TISSUE ANALYSIS Methods for cyanide determination in plant tissue have utilized various solvent extraction techniques with an emphasis on the determination of cyanide release potential, primarily from the breakdown of cyanogenic glycosides, rather than cyanide speciation and concentration [42–45]. The goal has been to determine the role of cyanide within naturally-occurring cycles as described in Chapter 3. As such, purification of the cyanogenic glycoside, rather than the removal of analytical interferences in the determination of cyanide species, is the objective. For cyanide species determination, some studies of plant tissue analysis have employed extraction methods similar to the conventional distillation [14], and microdiffusion [36] analytical techniques with colorimetric determination. Howe and Noble [46] digested plant tissue samples directly via distillation with MgCl2 and NaH2 PO4 prior to colorimetric development. Tittle et al. [47] acid-digested tissue via distillation with analysis of the liberated free

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cyanide both colorimetrically and by gas chromatography after bromination. Mizutani et al. [48] also utilized bromination for analysis of cyanide in apple samples, ground under distilled water. Bushey et al. [28] utilized a modified Standard Methods procedure involving homogenization with liquid nitrogen prior to extraction. Following homogenization, willow tissue samples, 1 to 1.5 g fresh-weight, were sonicated for 5 min, extracted in the dark for 16 h in 2.5 M NaOH, and analyzed without filtration for total and free cyanide by acid distillation and microdiffusion, respectively. Sample tissue extraction controls found recoveries of 89 and 100% for 100 ppb CNT as KCN and K4 Fe(CN)6 spiked in willow tissue slurries.

8.8.2 ANIMAL TISSUE SAMPLES Forensics analyses of cyanide in animal tissue samples have been performed by methods similar to that for aqueous samples. The analytical methods for total and free cyanide content of animal tissue are modified distillation [49] and microdiffusion [50] techniques. Analytical concerns for animal tissues are similar to those for plants in that the samples must be preserved to prevent cyanide release or species transformation prior to analysis. Also, the samples contain high amounts of various organics that have the potential to interfere with cyanide detection.

8.8.3 ALTERNATIVE EXTRACTION PROCEDURES FOR PLANT TISSUE Alternative solvents have been discussed for increasing cyanide recovery during the plant tissue extraction. Typical methods for stripping organic material from plant tissue involve a methanol/ chloroform (2:1 v/v) soak for three days [51,52]. Both solvents, but particularly methanol, have been used for the extraction of cyanogenic glycosides which are released to solution following extraction in polar solvents [43,44,53]. Hexane and 2-octanol have also been suggested for the cleanup of organic-rich liquid samples prior to analysis. However, methanol, hexane, and 2-octanol inclusion in the solvent matrix with 2.5 M NaOH interfered with the cyanide analytical technique while chloroform violently reacts with NaOH [54] and eventually free cyanide in solution [28]. Results obtained using only NaOH in the extraction medium avoided these interferences and provided good recoveries of cyanide spikes [28]. Also, extreme pH conditions, such as those used for preserving and analyzing aqueous cyanide samples, disable the functionality of the hydrolytic enzyme and prevent the artifact of cyanide synthesis during the analysis [55,56]. Filtration was shown to interfere with cyanide recovery while tissue did not interfere with the distillation or microdiffustion analysis [28]. Bushey et al. (unpublished results) have also found that sonication for more than five minutes or exposure to light during extraction, may reduce recovery of cyanide during the extraction process.

8.8.4 ANALYTICAL METHODS FOR USE WITH PLANT TISSUE A variety of analytical methods have been used to determine cyanide content of plant tissue extract slurry, including capturing released cyanide on picrate paper [57] according to the Feigl–Anger method [42], the use of NaOH-soaked paper [58], or the bromination of cyanide trapped in caustic solution [47,48]. Analysis via Picrate paper is subjective with lower precision, while the use of NaOH-soaked paper exhibited only 16% recovery [58]. Bromination involves additional sample handling, which increases the potential for free cyanide loss but does provide for increased accuracy and precision relative to the previous methods. The major limitation with respect to all three analyses is that only free cyanide is targeted unless the sample is acid digested and distilled prior to analysis. Extraction and analysis for total cyanide via distillation [14] and free cyanide via microdiffusion [36] provides cyanide speciation, in addition to total cyanide content [59].

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TABLE 8.1 Total Cyanide Recoveries from Spiked Soils using Alkaline Extraction and Analysis Procedure given in USEPA SW-846 Method 9013A

Sample Soil 1 Soil 2 Soil 3

Background total cyanide concentrationa (µg/g cyanide)

Spiking concentrationb (µg/g cyanide)

Average total cyanide recovery (%)b

Standard deviation (%)c

0.44 0.17 ND

500 500 500

95 78 86

1.8 (n = 2) 0.1 (n = 2) 6.4 (n = 4)

a Total cyanide was determined in the soil extraction solutions using USEPA SW-846 Method

9010B [19] and 9012A [61]. ND = nondetectable. b Soils were spiked using solid Fe [Fe(CN) ] prior to alkaline extraction specified in USEPA 4 6 3

SW-846 Method 9013A [18]. c n = Number of replicate analyses.

8.9 QUALITY CONTROL AND METHOD PERFORMANCE EVALUATION Method performance goals need to be established on a project-specific basis, as the type of solid or semi-solid sample and its physical–chemical characteristics, as well as the project context and goals, will determine the most appropriate approach to cyanide analysis [18,39]. The laboratory performing the analyses should establish quality control performance criteria for the application of any solid-specific methodology. USEPA guidance [60] on this topic provides a useful framework for developing such criteria. Some example performance data from USEPA for the application of the alkaline extraction technique described in Method 9013A [18] to three soils spiked with the solid Prussian Blue, Fe4 (Fe(CN)6 )3 (s), are presented in Table 8.1. The data are not recommended performance goals, but rather are examples from a single-laboratory study of what was achieved with the alkaline extraction technique with soils containing a common solid phase form of cyanide. As shown in Table 8.1, recoveries of 78 to 95% were attained with relative standard deviations ranging from 0.1 to 6.4%. These data illustrate the potential for high recovery of cyanide present in complex forms in soils, but as has been emphasized in this chapter, different extraction and analysis approaches may be needed to achieve similar recoveries with other solids or semi-solids.

8.10 SUMMARY AND CONCLUSIONS • Cyanide can be associated with solids in contaminated water and soil systems in a number of ways: as solid-phase cyanide compounds including metal-cyanide compounds and, to a lesser extent, as simple cyanide salts; as adsorbed species on solid surfaces; and as dissolved, aqueous phase cyanide species via entrapment of contaminated water in pore spaces. • Analysis of cyanide in solids is usually performed by extracting the cyanide into aqueous solution with strong acid, strong base, or reagent water, separating solid and water at the end of a designated extraction period, and then analyzing the extract for total cyanide or other cyanide species using aqueous phase analysis techniques. • Direct acid distillation and analysis of solid samples can also be performed, but in many cases will not yield complete recovery of cyanide.

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• The different extraction solutions can yield substantially different results for a particular solid, as pH has a very strong effect on cyanide concentrations in the leachate. • Adjustment of pH of the solid phase alkaline extract solutions is essential prior to subjecting the solution to free cyanide analysis by microdiffusion or metal–cyanide complex analysis by ion chromatography. • For many cyanide-contaminated solids, especially those with iron-cyanide solids dominant, extraction performed under alkaline (pH > 12) conditions with high LS ratio results in maximum cyanide recovery. • The LS ratio employed in the extractions typically ranges from about 20:1 w/w to 1000:1 w/w; extraction yields depend on LS ratio for many solids. • For a solid sample of unknown cyanide content and composition, it is advisable to use extract solutions spanning low to high pH, as well as a range of LS ratios, depending on the goals of the analysis. • The extraction time required for recovery of cyanide from a solid sample depends on the extractant solution, the cyanide analyte of interest, and the sample matrix. • Cyanide analysis in high-water-content, high-organic-content solids such as plant tissue, animal tissue, and sludges requires special considerations to avoid analytical artifacts.

REFERENCES 1. Ford-Smith, M.H., The Chemistry of Complex Cyanides: A Literature Survey, Department of Scientific and Industrial Research, National Chemical Laboratory, London, 1964. 2. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 3. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 4. Theis, T.L. and West, M.L., Effects of cyanide complexation on the adsorption of trace metals at the surface of goethite, Environ. Technol. Lett., 7, 309, 1986. 5. Rennert, T. and Mansfeldt, T., Sorption of iron–cyanide complexes in soils, Soil Sci. Soc. Am. J., 66, 437, 2002. 6. Chatwin, T.D., Zhang, J., and Gridley, G.M., Natural mechanisms in soil to mitigate cyanide release, in Proc. Superfund ’88, The 9th National Conference, Hazardous Materials Control Research Institute, Washington, DC, 1988, p. 467. 7. Rennert, T. and Mansfeldt, T., Sorption of iron-cyanide complexes on goethite in the presence of sulfate and desorption with phosphate and chloride, J. Environ. Qual., 31, 745, 2002. 8. Bushey, J.T. and Dzombak, D.A., Ferrocyanide adsorption on aluminum oxides, J. Coll. Int. Sci., 272, 46, 2004. 9. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sediment Contamination, submitted, 2005. 10. Keith, L.H., Ed., Principles of Environmental Sampling, American Chemical Society, Washington, DC, 1988. 11. USEPA, Methods for evaluating the attainment of cleanup standards. Vol. 1, Soils and solid media, EPA-230/02-89-042, U.S. Environmental Protection Agency, Office of Policy, Planning and Evaluation, Washington, DC, 1989. 12. USEPA, Subsurface characterization and monitoring techniques, Vol. II. The vadose zone, field screening and analytical methods, EPA-625/R-93-003b, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 1993. 13. USEPA, Subsurface characterization and monitoring techniques, Vol. I. Solids and ground water, EPA-625/R-93-003a, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 1993. 14. APHA/AWWA/WEF, Method 4500-CN: Cyanide, in Standard Methods for the Examination of Water and Wastewater, 20th ed., L.S. Clesceri, A.E. Greenberg, and A.D. Eaton, Eds., American Public

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Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. USGS, Method I-6302-85: Cyanide, recoverable-from-bottom-material; colorimetric, barbituric acid, automated-segmented flow, USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, http://www.nemi.gov, 1985. USGS, Method I-5300-85: Cyanide, total-in-bottom-material; colorimetric, pyridine-pyrazolone, USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, http://www.nemi.gov, 1985. APHA/AWWA/WEF, Method 207: Cyanide, in Standard Methods for the Examination of Water and Wastewater, 13th ed., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1971. USEPA, Method 9013A: Cyanide extraction procedure for solids and oils, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 1, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 2004. USEPA, Method 9010B: Total and amenable cyanide: Distillation, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 5, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. Mansfeldt, T. and Biernath, H., Determination of total cyanide in soils by micro-distillation, Analyt. Chim. Acta, 406, 283, 2000. Theis, T.L., Young, T.C., Huang, M., and Knutsen, K.C., Leachate characteristics and composition of cyanide-bearing wastes from manufactured gas plants, Environ. Sci. Technol., 28, 99, 1994. Shifrin, N.S., Beck, B.D., Gauthier, T.D., Chapnick, S.D., and Goodman, G., Chemistry, toxicology, and human health risk of cyanide compounds in soils at former manufactured gas plant sites, Regul. Toxicol. Pharmacol., 23, 106, 1996. USEPA, Method 1311: Toxicity characteristic leaching procedure, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 5, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. ASTM, ASTM D5233-92 Standard test method for single batch extraction method for wastes, in Annual Book of ASTM Standards, Vol. 11.04, ASTM International, West Conshohocken, PA, 1999. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Dissolution behavior of iron cyanide (Prussian Blue) in contaminated soils, Environ. Sci. Technol., 26, 1832, 1992. USEPA, Method 1312: Synthetic precipitation leaching procedure, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 5, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. ASTM, ASTM D3987-85. Standard test method for shake extraction of solid waste with water, in Annual Book of ASTM Standards, Vol. 11.04, ASTM International, West Conshochoken, PA, 2004. Bushey, J.T., Ebbs, S.D., and Dzombak, D.A., Plant tissue extraction method for complexed and free cyanide, Water, Air and Soil Pollut., 157, 281, 2004. Mansfeldt, T. and Biernath, H., Method comparison for the determination of total cyanide in deposited blast furnace sludge, Analytica Chimica Acta, 435, 377, 2001. Mansfeldt, T., Leyer, H., Barmettler, K., and Kretzschmar, R., Cyanide leaching from soil developed from coking plant purifier waste as influenced by citrate, Vadose Zone J., 3, 471, 2004. Ghosh, R.S., Nakles, D.V., Murarka, I., and Neuhauser, E.F., Cyanide speciation in soil and groundwater at manufactured gas plant (MGP) sites, Environ. Eng. Sci., 21, 752, 2004. Köster, H.W., Risk assessment of historical soil contamination with cyanides: origin, potential human exposure and evaluation of intervention values, RIVM report 711701019, Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment), Bilthoven, The Netherlands, 2001. Mansfeldt, T., Gehrt, S.B., and Friedl, J., Cyanides in a soil of a former coking plant site, Z. Pflanzenernähr. Bodenk., 161, 229, 1998. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Solubility of cyanide in contaminated soils, J. Environ. Qual., 23, 785, 1994. Young, T.C. and Theis, T.L., Determination of cyanide in manufactured gas plant purifier wastes, Environ. Technol., 12, 1063, 1991.

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36. ASTM, Designation D 4282-95. Standard test method for determination of free cyanide in water and wastewater by microdiffusion, in Annual Book of ASTM Standards, Vol. 14.02, ASTM International, West Conshohocken, PA, 1998, p. 468. 37. Ghosh, R.S., Drop, S., and Smith, J.R. Performance of anion exchange chromatography method for the routine evaluation of metal cyanide complexes in solid waste leachates, in Proc. 20th Annual National Environmental Monitoring Conference, U.S. Environmental Protection Agency, Washington, DC, 2004, p. 156. 38. USEPA, Method 9015: Metal cyanide complexes by anion exchange chromatography and UV detection, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 0, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 2004. 39. Kosson, D.S., van der Sloot, H.A., Sanchez, F., and Garrabrants, A.C., An integrated framework for evaluating leaching in waste management and utilization of secondary materials, Environ. Eng. Sci., 19, 159, 2002. 40. USEPA, Method 335.2 (CLP-M): Cyanide, total (titrimetric; spectrophotometric) via midi-distillation, U.S. Environmental Protection Agency, Superfund Analytical Services, Contract Laboratory Program, Washington, DC, http://www.epa.gov/superfund/programs/clp/ilm5.htm, accessed: July 2, 2004. 41. Lupichuk, W., Theis, T.L., and Young, T.C., Analysis of cyanide species in aluminum potlining wastes and leachates, Report ETC 95-10-03, prepared by Clarkson University for Alcoa Technical Center, Alcoa Center, PA, 1995. 42. Aikman, K., Bergman, D., Ebinger, J., and Seigler, D., Variation of cyanogenesis in some plant species of the midwestern United States, Biochemical Systematics and Ecology, 24, 637, 1996. 43. Kobaisy, M., Oomah, B.D., and Mazza, G., Determination of cyanogenic glycosides in flaxseed by barbituric acid-pyridine, pyridine-pyrazolone, and high-performance liquid chromatography methods, J. Agric. Food Chem., 44, 3178, 1996. 44. Selmar, D., Grocholewski, S., and Seigler, D.S., Cyanogenic lipids, Plant Physiol., 93, 631, 1990. 45. Selmar, D., The cleavage of cyanogenic lipids by esterases, Physiologia Pantarum, 83, 63, 1991. 46. Howe, M. and Noble, D., Effect of cyanide residue on vegetation bordering a Black Hills stream, Proc. S.D. Acad. Sci., 64, 112, 1985. 47. Tittle, F.L., Goudey, J.S., and Spencer, M.S., Effect of 2,4-dichlorophenoxyacetic acid on endogenous cyanide, β-cyanoalanine synthase activity, and ethylene evolution in seedlings of soybean and barley, Plant Physiol., 94, 1143, 1990. 48. Mizutani, F., Hirota, R., and Kadoya, K., Cyanide metabolism linked with ethylene biosynthesis in ripening apple fruit, J. Japan. Soc. Hort. Sci., 56, 31, 1987. 49. Nolte, K.B. and Dasgupta, A., Prevention of occupational cyanide exposure in autopsy prosectors, J. Forensic Sci., 41, 146, 1996. 50. Swanson, J.R. and Krasselt, W.G., An acetonitrile-related death, J. Forensic Sci., 39, 271, 1994. 51. Cohen, C.K., Fox, T.C., Garvin, D.F., and Kocian, L.V., The role of iron-deficiency stress responses in stimulating heavy metal transport in plants, Plant Physiol., 116, 1063, 1998. 52. Hart, J.J., Norvell, W.J., Welch, R.M., Sullivan, L.A., and Kocian, L.V., Characterization of zinc uptake, binding, and translocation in intact seedlings of bread and durum wheat cultivars, Plant Physiol., 118, 219, 1998. 53. Forslund, K. and Jonsson, L., Cyanogenic glycosides and their metabolic enzymes in barley, in relation to nitrogen levels, Physiologia Pantarum, 101, 367, 1997. 54. McKetta, J.J. and Cunningham, W.A., eds., Encyclopedia of Chemical Processing and Design, Marcel Dekker, New York, NY, 1976. 55. Halkier, B.A. and Moller, B.L., The biosynthesis of cyanogenic glucosides in higher plants, J. Biol. Chem., 265, 21114, 1990. 56. Lechtenberg, M., Nahrstedt, A., Wray, V., and Fronczek, F.R., Cyanoglucosides from Osmaronia Cerasiformis (Rosaceae), Phytochemistry, 37, 1039, 1994. 57. Jacobs, K.A., Santamour, F.S., Johnson, G.R., and Dirr, M.A., Differential resistance to entomosporium leafspot disease and hydrogen cyanide potential in photinia, J. Environ. Tech., 14, 154, 1996. 58. Grossman, K. and Kwiatkowski, J., Evidence for a causative role of cyanide, derived from ethylene biosynthesis, in the herbicidal mode of action of quinclorac in barnyard grass, Pesticide Biochem. Physiol., 51, 150, 1995.

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59. Ebbs, S.D., Bushey, J.T., Poston, S., Kosma, D., Samiotakis, M., and Dzombak, D.A., Transport and metabolism of free cyanide and iron cyanide complexes by willow, Plant Cell Env., 26, 1467, 2003. 60. USEPA, Quality control, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 1, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1992. 61. USEPA, Method 9012A: Total and amenable cyanide (automated colorimetric, with off-line distillation), Rev 1, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1996.

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and Transport of 9 Fate Anthropogenic Cyanide in Surface Water

Thomas C. Young, Xiuying Zhao, and Thomas L. Theis CONTENTS 9.1 9.2

Synoptic Review of Cyanide Speciation and Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Primary Cyanide Fate-Determining Environmental Processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.1 Photolytic Reactions Involving Complexed Cyanides. . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.1.1 Photolytic Process Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.1.2 Photolytic Process Modeling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.2 Air–Water Exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.2.1 Literature Observations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.2.2 Estimating Mass Transfer from Two-Film Theory . . . . . . . . . . . . . . . . . . 9.2.3 Biological Fate-Determining Processes for Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3 Prediction and Assessment of Exposure to Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.1 Modeling Framework for Natural Aquatic Systems — The Mass Balance Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.2 Case Study: Fate of Cyanide in a Small Stream Environment . . . . . . . . . . . . . . . . . . 9.3.2.1 Calibration Approach — Average Field Conditions . . . . . . . . . . . . . . . . 9.3.2.2 Calibration Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

172 174 174 174 176 176 176 177 178 179 180 182 184 185 188 188

The environmental toxicity of cyanide has been widely recognized, and the occurrence of accidental, but environmentally catastrophic, spills of impounded, cyanide-containing waters has engendered widespread concern. Two widely publicized examples of cyanide pollution of natural freshwaters include the Summitville gold mining operation (Colorado, The United States) where pollution of the Alamosa River began during the early 1990s [1]; and the Aurul mining operation at Baia Mare (Romania), where a spill occurred, beginning in January 2000, that severely affected the Somes–Tisza–Danube River ecosystem before reaching the Black Sea [2]. It is as true for cyanide as it is for other toxic chemicals that the risks associated with release into a receiving water largely depend on site- and time-specific environmental conditions. These dependencies inevitably arise because the situation-specific conditions exert a major influence over the numerous physical and chemical interactions that determine environmental fate. Examples of environmental conditions that affect the exposure risk associated with a release of cyanide to an aquatic system include concentrations and pH-dependent speciation of cyanide and metallic trace elements (especially iron, Fe), the availability and quality of solar radiation, the nature of the biotic 171

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community of the receiving water, fluid flow characteristics, and the temperature of both air and water. This chapter addresses the environmental transport and fate of cyanide species commonly encountered in contaminated freshwater aquatic systems by: (i) briefly considering the principal environmental processes that determine the toxicity and fate of cyanide in aquatic systems under typical environmental conditions; and (ii) the development and case-study implementation of a modeling framework to facilitate assessment of cyanide fate and exposure in an affected natural aquatic system. The processes included in the model presented herein include photolytic decay of complex cyanide species (the commonly occurring species, Fe(CN)6 −4 or hexacyanoferrate (+II), was selected as a model compound), biological uptake/utilization/decay/transformation of HCN, and water– air mass transfer of HCN.

9.1 SYNOPTIC REVIEW OF CYANIDE SPECIATION AND TOXICITY Inorganic compounds that contain the cyanide/cyanogen group (C≡N− ) exhibit a wide diversity of physical–chemical behavior. According to Ford-Smith [3] on the order of 28 elements, for a total of 64 different element–oxidation state combinations, may form complex cyanides. To facilitate discussion, a terminology will be employed that, with few changes, follows that used by Doudoroff [4], and later by Hartung [5], for cyanide forms in freshwater aquatic systems (Table 9.1). The listing of cyano species in Table 9.1 is very simple and far from exhaustive ( [6–8]), but serves to focus the discussion on species that have principal significance in the areas of environmental impact and toxicological effects. As discussed in Chapter 14, it has been well established that toxicity to aquatic biota associated with exposure to a cyanide-containing waste or pollutant derives from the nascent or actual exposure to free cyanide in the waste. Exposure of aquatic biota to complex cyanide, on the other hand, can be virtually nontoxic, because the toxicity of complex cyanide nearly always is due to the presence of free cyanide formed by dissociation of the metal complex. Few exceptions to this rule have been observed, and those that have been reported note that the apparent toxicity of the complex cyanide actually has other causes. For example, during fish bioassay studies involving two

TABLE 9.1 General Descriptions of Cyano-Species of Interest in Aquatic Systems Name

Description

Ionic cyanide Molecular cyanide Free cyanide Simple cyanide Weak acid dissociable cyanide

Complex cyanide

Cyanide ion, CN− Hydrocyanic acid, HCN HCN + CN− Compound that dissociates readily in water to form a single cation (e.g., K+ ) and CN− anion Analytically-defined form consisting of Free CN + Simple CN + Complex CN (see below) with low resistance to hydrolysis under mildly acidic conditions (e.g., tetracyanonickelate, (Ni(CN)−2 4 ), tetracyanozincate, −2 (Zn(CN)−2 4 ), tetracyanocuprate, (Cu(CN)4 ) Compound that dissociates in H2 O to yield one or more cations (e.g., K+ ) and one or more cyanogen groups coordinated with a central cation to −3 form an anion (e.g., ferrocyanide: Fe(CN)−4 6 , ferricyanide: Fe(CN)6 ); complexes vary in resistance to mild acid hydrolysis

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metal-cyanide complexes less frequently encountered in natural waters, Doudoroff [9] determined that free Cu ion, rather than Cu-complexed cyanide, was responsible for the toxicity associated −2 with Cu(CN)− 2 /Cu(CN)3 in aqueous solutions of these chemical species. These observations were confirmed by Broderius [10], who also found that free Ag+ displayed a similar enhancement of the −4 apparent toxicity of Ag(CN)− 2 . In the specific instance of the hexacyanoferrate complexes, Fe(CN)6 and Fe(CN)−3 6 , which are very commonly encountered cyanide complexes in surface waters, the results of Doudoroff and coworkers [4,9,11,12], Broderius [10], and others have repeatedly demonstrated that the hexacyanoferrate complexes are essentially nontoxic, and that such toxicity that may be associated with these complexes is due to dissociation of the complexes to yield free cyanide. The foregoing brief discussion of cyanide toxicity in aquatic environments receives a complete and comprehensive treatment in Chapter 14. In spite of its limited scope at this juncture, however, the cited evidence clearly supports the view that in order to be useful for design and management applications, a cyanide model must focus on the main processes and reactions that describe the sources, transport, and fate of free cyanide, including significant interactions with other cyanidecontaining materials. As reviewed and discussed in some detail in the next sections of this chapter, the principal processes of interest include photolytic transformation of hexacyanoferrate to free cyanide, plus volatilization and biotic transformation and utilization of free cyanide. The relationship between these processes and transport in a context that supports model development is illustrated schematically in Figure 9.1. The focus of the remainder of this chapter includes the development of a modeling framework for the environmental fate of cyanide that incorporates the features shown in Figure 9.1, and provides an application of the modeling framework to a case study involving the ongoing contamination of receiving waters by cyanide-bearing wastes. It should be noted that under special circumstances, as discussed in other chapters of this text, other processes may also influence the environmental fate of cyanide. Examples of these include hydrolysis, adsorption, precipitation, photooxidation, and sediment–water interactions. Furthermore, the rate and extent of any of these processes can be influenced by such site-specific environmental variables as pH, pE, the concentrations of complexing cations and competing ligands, and temperature. These details and complexities, for the most part, are not addressed in the model development and example application presented here.

Atmosphere Water Column

Volatilization

External Loading Free and HydrolyticallyLabile Cyanides, “HCN”

Advection

Dispersion

Photolysis Complex, Photolyzable Cyanides, “FeCN”

Advection

Dispersion Biotic Transformation and Utilization

FIGURE 9.1 Schematic diagram of the primary transport and fate processes that affect cyanide in natural surface waters.

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9.2 PRIMARY CYANIDE FATE-DETERMINING ENVIRONMENTAL PROCESSES 9.2.1 PHOTOLYTIC REACTIONS INVOLVING COMPLEXED CYANIDES 9.2.1.1 Photolytic Process Characterization Several cyanometallate complexes are known to display photochemical reactivity [13–15], including the cyanide complexes involving Mo(+IV) and Mo(+V), W(+IV) and W(+V), Cr(+III), Mn(+III), Fe(+II) and Fe(+III), and Co(+III). Nevertheless, it is only the hexacyanoferrate com−3 plexes, Fe(CN)−4 6 and Fe(CN)6 , that are encountered with sufficient frequency and at sufficient concentrations in surface waters, to be of concern for managing water quality in most systems of interest. The photolytic decay reactions of the hexacyanoferrates, however, comprise a set of reactions that have generated a variety of interests for many years. In 1901, Matuschek (as cited in Foster [16]) noted that Fe(OH)3 precipitated from a solution of K4 Fe(CN)6 when the latter was exposed to sunlight, and that KCN was always present in the solution afterward. Other early workers demonstrated that UV light energy had the ability to convert Fe(CN)−4 6 to Fe(OH)3 and KCN in an aqueous environment. Foster [16] noted the effect of UV light and interpreted his observations as suggesting the possible formation of Fe(CN)−3 6 , and subsequent hydrolysis of the latter to yield KCN. Schwartz and Tede [17], however, hypothesized a photoaquation reaction involving K4 Fe(CN)6 that was accompanied by the formation of free cyanide in aqueous solution. −3 It is now known that the primary reaction for the photodecomposition of Fe(CN)−4 6 and Fe(CN)6 , shown as Equations (9.1) and (9.2), has been demonstrated by Ašpergˇer [18], Mitra et al. [19], and MacDiarmid and Hall [14]. −3 Fe(CN)−4 + HCN + OH− 6 + 2H2 O ←→ Fe(CN)5 · (H2 O)

(9.1)

Fe(CN)−3 6

(9.2)

+ 2H2 O ←→ Fe(CN)5 · (H2 O)

−2

+ HCN + OH

−

Moreover, Ašpergˇer [18] and Balzani and Carassiti [20] have indicated that the pentacyanoferrate product of the photoaquation reaction, Fe(CN)5 · (H2 O)−3 , upon prolonged UV light exposure, can undergo sequential photosubstitution reactions to release all the cyanide from the iron complex, and indeed, complete photolytic dissociation has been observed routinely under environmental conditions [21]. Other information about the photolysis of the hexacyanoferrate species that is known with reasonable certainty includes: • UV radiation from the sun or other environmental source, in the presence of a suitable −3 electron acceptor, will permit Fe(CN)−4 6 to photooxidize to yield Fe(CN)6 , as illustrated in Equation (9.3) [22]. − Fe(CN)−4 −→ Fe(CN)−3 6 6 +e . hv

(9.3)

• Shirom and Stein [23,24] found that significant electron production (hence, oxidation of Fe(+II) to Fe(+III)) by photon absorption occurred only at λ < 313 nm, though photoaquation was enhanced at longer wavelengths, 313 < λ < 365 nm. • MacDiarmid and Hall [14] found that both hexacyanoferrate complexes underwent photolysis in full sunlight (as measured by pH changes attributed to cyanide hydrolysis), and the photolysis of both complexes was at least partially reversible (dark reformation −3 and CN− and Fe(CN)−3 from Fe(CN) · (H O)−2 of Fe(CN)−4 2 5 6 from Fe(CN)5 · (H2 O) 6 − and CN ).

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• Gáspár and Beck [25] proposed and calibrated a mathematical model for the proposed reaction kinetics, shown here as Equation (9.4), where φ is the quantum yield, Ic is the photon absorption rate of the complex, and k6 , k−6 are reaction rate coefficients. −

d{[Fe(CN)6 ]−4 } = φIc + k6 {[Fe(CN)6 ]−4 } − k−6 {[Fe(CN)5 H2 O]−3 } [CN] dt

(9.4)

Much of the literature on photolysis of hexacyanoferrates has pertained to experimentation that has been highly controlled and aimed at elucidating reaction mechanisms and kinetics. Clearly, this is needed for developing the deepest understanding of the environmental significance of these substances in surface waters. Still, the complexity of natural waters will continue to present a difficult problem in translation, beyond the control of the laboratory. Important advances in understanding the photolysis of hexacyanoferrate in natural waters were achieved by Broderius and Smith [26] through controlled experimental manipulation, in vitro. The general approach used by Broderius and Smith [26] included monitoring buffered solutions (pH 7.8) of varying initial concentrations of K4 Fe(CN)6 and K3 Fe(CN)6 , either in open vessels (battery jars) or head space-free vials (“filled to the brim”), during thermostated (20◦ C) water bath exposure to sunlight. Free cyanide concentration time series were analyzed to estimate photolysis rate constants and half-lives (first-order photolytic reaction of the hexacyanoferrates was assumed). Recent information, however, demonstrates significant differences exist between the relative rates − of Fe(CN)−4 6 photolysis and CN formation [21], an observation which is consistent with the photolysis model proposed by Gáspár and Beck [25]. Nevertheless, the results and conclusions drawn from these experiments with natural waters were wide-ranging, and have significantly influenced the current regulatory environment. The main points of the information gathered, and the conclusions made by Broderius and Smith [26], included the following items: • Dissolved oxygen concentrations, typical of surface waters, had no effect on photolysis rates, while pH had only a minor effect on the rates, which increased on the order of 15 to 30% as pH decreased from 9 to 6.6. • The maximum amount of free cyanide that could be released from prolonged photochem−3 ical reaction of Fe(CN)−4 6 and Fe(CN)6 amounted to 85 and 49%, respectively. This reflects a stoichiometry of releasing 5/6 and 3/6 of the total cyanide initially present in the unreacted complexes, respectively. Such a finding is not consistent with any particular photolytic reaction model, and others have observed virtually 100% conversion of Fe(CN)−4 6 to products. For example, Kuhn and Young [21] repeatedly observed complete −1 conversion within 30 min of exposure of low concentrations of Fe(CN)−4 6 (≤400 µg L as CN) to full sunlight. • Using a method described by Zepp and Cline [27], a curve fit was performed to estimate −3 the quantum yield of the photolysis reactions. For Fe(CN)−4 6 and Fe(CN)6 , the values were 0.14 and 0.0023, respectively. • Light attenuation vertically, through a turbid suspension of particulate matter, affected the distribution of photolysis in surface waters. Photolysis rates declined exponentially with depth, as would be predicted from theory, and the photolysis rate for Fe(CN)−4 6 decreased — also as expected, due to the shorter more rapidly with depth than that for Fe(CN)−3 6 excitation wavelengths of the former complex. The results also displayed, however, a marked lack of reaction attenuation with concentration of suspended particulate matter. Kuhn and Young [21] made similar observations about the effect of small amounts of kaolin clay on the rate of Fe(CN)−4 6 photolysis. • Nearly equal photolysis rates were observed for Fe(CN)−4 6 in solutions prepared from several sources of natural water, plus buffered deionized water. None of the natural waters

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tested by Broderius and Smith [26] substantially enhanced or inhibited free cyanide formation from the Fe(II) complex. Kuhn and Young [21] confirmed these observations for two surface waters from northern New York having varying color and turbidity, but Kuhn and Young also noted a seemingly significant reduction in the rate of photolytic decay for Fe(CN)−4 6 present in contaminated groundwater, compared to the two surface waters. 9.2.1.2 Photolytic Process Modeling The kinetic model of hexacyanoferrate photolysis presented previously in Equation (9.4) is best used to advance understanding of photolytic reaction mechanisms, but a different approach is required to model the photolytic decay of these compounds in surface waters. An especially useful method is termed the “near surface approach,” and permits estimation of an integrated average rate of photolysis that applies to the entire photoactive depth of the system of interest [28]. The near-surface approach yields an estimate of the first-order photolysis rate coefficient, kp , according to Equation (9.5):

kp = kp0

ID 1 − exp[−αD (λ∗ )Z] I 0 D0 αD (λ∗ )Z

(9.5)

where kp0 is the first-order, direct photolysis rate coefficient, measured at the surface of the water body (Z = 0), [d−1 ]; I0 , I the total solar radiation at the surface of the water body (Z = 0) and at depth Z, [ly d−1 ]; D, D0 the system-specific, dimensionless constants determined by the radiance distribution function, magnitude approximately 1.33; and αD (λ∗ ) the extinction coefficient at wavelength of maximum spectral absorbance for surface water, dependent on light absorption by water, chlorophyll, dissolved organic carbon, and suspended solids [m−1 ]. Constraints on the validity of Equation (9.5) for estimating direct photolysis rates in natural systems are noted in Mills et al. [29].

9.2.2 AIR–WATER EXCHANGE 9.2.2.1 Literature Observations Although the hexacyanoferrates, as dissolved salts, have no vapor pressure of consequence (the solid sodium and potassium salts decompose when heated sufficiently in vacuo [6]), the hydrocyanic acid fraction of free cyanide has a significant vapor pressure, 53.1 kPa at 10.2◦ C [30]. Leduc et al. [30] provide a value for the distribution of HCN between air and water: 3 × 10−3 mg HCN L−1 (air)/mg HCN L−1 (water). However, the pH and temperature dependence of this distribution coefficient was not given [31]. Additional information on equilibrium air–water partitioning of HCN is provided in Chapter 5. Broderius and Smith [26] measured loss rates of HCN from open containers, but did not attempt to quantify the Henry’s Law constant. Indeed, their results did not follow an apparent first-order mass transfer function with respect to HCN concentration, though the experimental protocol was not intended to be so precise as to yield that information. Volatilization of HCN from open containers was slow, and exhibited half-lives (t1/2 ) from ca. 24 to 60 h for 25 to 10◦ C, respectively. The results were inconsistent, with a first-order model for the lower temperatures, however, and volatilization rates in the laboratory were about half of those observed during rooftop exposures, a difference the authors attributed to wind effects. Agitation was infrequent and not controlled. In assessing the significance of these observations, the authors stated their belief that volatilization losses were slower than the formation of HCN would be in surface waters from photolysis of hexacyanoferrate complexes, though rate data given later in the report suggested otherwise.

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9.2.2.2 Estimating Mass Transfer from Two-Film Theory HCN mass transfer between water and air can be estimated from a combination of considerations, including classical two-film theory [32] and adjustments for varying environmental conditions, such as temperature, hydrodynamics, and wind-generated turbulence. Using this approach, the overall liquid-side mass transfer coefficient may be estimated by considering mass transfer resistances through both the liquid and gas phases, as expressed in Equation (9.6). 1 RT 1 = + KOL KL HHCN ∗ Kg

(9.6)

where KOL is the overall liquid-side mass transfer coefficient, [m d−1 ], KL the liquid phase mass transfer coefficient, [m d−1 ], Kg the gas phase mass transfer coefficient, [m d−1 ], HHCN the Henry’s law constant for equilibrium partitioning of HCN between the liquid and gas phases, [m3 atm mol−1 ], R the universal gas constant, [8.206 × 10−5 atm m3 (mol K)−1 ], and T the absolute temperature, [K]. The overall mass transfer coefficient of a chemical generally depends on both the liquid- and gasphase resistances, though the resistance on one side may predominate to limit the overall mass transfer rate. Liquid-side mass transfer would control volatilization if the Henry’s law constant for HCN (HHCN ) exceeded 4.6 atm L mol−1 , while the gas-phase resistance would control volatilization if HHCN were less than 0.013 atm L mol−1 . Resistance from both phases would contribute significantly to the overall resistance for intermediate values of HHCN , and, indeed, that is the case for HCN, as the values of Henry’s law constant for HCN have been reported in the range from 0.073 to 0.122 atm L mol−1 (see Chapter 5). Bodek et al. [33] described the temperature dependence of the Henry’s Law constant for HCN as shown in Equation (9.7) (HHCN in mmHg L mol−1 , temperature in K): log HHCN = −

1272.9 + 6.238 T

(9.7)

while Yoo et al. [34] developed the relationship shown in Equation (9.8) (HHCN in atm L mol−1 , temperature in K): ln H = 9.5850 − 0.03147 T + 3.1704 ln T − 6302.0/T

(9.8)

The rates of liquid- and gas-film mass transfer are rarely measured for HCN and published (see, however, Ref. [35]). They may be estimated, nevertheless, by correlation methods, using mass transfer rates determined for other, better-characterized substances. The mass transfer characteristics of oxygen have been studied widely, and provide a convenient reference compound to estimate the liquid-film mass transfer coefficient (KL ) for many other chemicals using the relationship shown in Equation (9.9) [28,29]:
KL,HCN = KL,O

32 MHCN

0.25 (9.9)

where KL,HCN is the liquid-film mass transfer coefficient for HCN, MHCN the molecular weight of HCN, and KL,O the oxygen-transfer coefficient in the water phase. Similarly, the gas-film mass transfer coefficient (KG ) can be estimated from water vapor transfer rate in air, as shown

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in Equation (9.10):
KG,HCN = KG,H2O

18 MHCN

0.25 (9.10)

where KG,HCN is the gas-film mass transfer coefficient and KG,H2O is the water vapor transfer rate in air [28,29]. Liss and Slater [36] suggested typical environmental values of KL and KG to be 0.2 m h−1 for oxygen transfer and 30 m h−1 for water vapor transfer, respectively. Many and varied formulations have been proffered to predict oxygen reaeration in streams and rivers. Three of these are used most commonly: O’Connor–Dobbins, Churchill, and Owens–Gibbs [28], all of which were developed based on the surface-renewal model for different types of streams, using experimental or empirical approaches. Among these three, the Owen–Gibbs relationship is best applied to shallow systems, which describe conditions that apply to the illustrative case study that follows. The Owen–Gibbs relationship can be described as shown in Equation (9.11): Ka = 5.32

U 0.67 Z 1.85

so,

U 0.67 = 5.32 0.85 Z

KL,O

Ka =

and

KL,O Z

(9.11)

where Ka is the oxygen reaeration rate coefficient, d−1 , U is water velocity, m s−1 , and Z is water depth, m. Similarly, the gas-film coefficient for water (KG,H2O , m d−1 ) is related to wind speed (Uw , m s−1 ) and, using Equation (9.10), can be approximated by Equation (9.12) [28,29]: KG,H2O = 168Uw KG,HCN = 168Uw




18 MHCN

0.25

(9.12)

In a study of a small stream contaminated with cyanide in eastern Tennessee, Kavanaugh et al. [35] found reasonably good agreement between HCN mass transfer rates measured in situ, using flux chambers, and values computed using two-film theory that incorporated modifications for temperature and wind speed. Values for the overall rate coefficient for liquid mass transfer for HCN measured in situ ranged from 1.2 to 42.2 × 10−3 m d−1 , while values for the computed value for KL,HCN ranged from 14.9 to 33 × 10−3 m d−1 in the two systems. The latter values are well within the range of observed values, which suggests the fitness of the approach provided by the modified two-film theory for HCN mass transfer rate estimation for modeling purposes.

9.2.3 BIOLOGICAL FATE-DETERMINING PROCESSES FOR CYANIDE A comprehensive examination of the known biological reactivity of cyanide species may be found in Chapter 6 and elsewhere [35,37,38]. The information contained in Chapter 6 indicates that generally, free cyanide readily biodegrades, as long as the HCN exposure concentration does not exceed toxicity limits, and the ease of biodegradation for a given biotic community will vary in expected ways, depending on the specified array of environmental conditions (e.g., temperature, dissolved oxygen levels, mixing and turbulence, nutrients, form and abundance of alternative nitrogen sources, and more). Moreover, experimental or other data that suggests biodecay, biouptake, or −4 biotransformation of Fe(CN)−3 6 , Fe(CN)6 , or other cyanometallate complexes is lacking. Hence, for modeling purposes, it may be assumed that free HCN is the bioavailable, bioactive form of cyanide.

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Depending on the nature of the biotic community, degradation/utilization pathways will be represented by a mixture of quasi-homogeneous reactions, for example, uptake and decay of free cyanide brought about by some combination of biota in the sestonic community, such as macro- and microbiota in suspended culture or other “dispersed phase” growth; and, mixed heterogeneous reactions, for example, uptake and decay of cyanide species brought about by some combination of biota in the periphyton, epipellic, and epilithic communities, or possibly direct uptake/utilization by rooted vascular hydrophytes. In the latter case, environmental conditions may demand considering uptake of free cyanide from sediment interstitial water, and possibly, the effects of mycorrhizal interactions on free cyanide fate. The simplest approach to addressing the complex set of factors that influence the biodecay of cyanide species in surface water, in a way that would be adequate to satisfy a wide range of modeling objectives, would be to assume that only HCN is subject to biodecay/biouptake. Then the aggregate effect of all such reactions can be represented by a first-order homogeneous reaction process, using an undifferentiated decay coefficient, kb , as illustrated in Equation (9.13): dCHCN = −kb CHCN dt

(9.13)

In taking this simple first-order approach, it is important to recognize that the decay coefficient, kb , is highly system-specific, and will vary both temporally and spatially due to fluctuations and trends in such process-determining environmental factors as temperature, mixing, inter-species competition and biodiversity, and others. Moreover, the nature of biochemical decay and utilization may not be first-order, but may be better represented by a mixed-order model, such as that represented by the Monod Equation, or any of several other nonlinear representations [39]. Further work to characterize cyanide utilization will be required, to determine the most appropriate biokinetic representation. Finally, environmentally-appropriate values for the decay coefficient, kb , cannot be easily acquired through controlled experimentation in vitro; however, they may be estimated satisfactorily for site-specific applications by model calibration procedures.

9.3 PREDICTION AND ASSESSMENT OF EXPOSURE TO CYANIDE Numerical simulation models provide a means to investigate the effects of multiple, simultaneously interacting processes that determine contaminant exposure in receiving waters. The information discussed in the earlier part of this chapter, therefore, provided the basis for developing a numerical simulation framework to evaluate the effects of fate-determining processes on cyanide exposure in freshwater aquatic systems. As illustrated in Figure 9.1 and discussed previously, the modeling activity reported here addresses the transport and fate of hexacyanoferrate and free cyanide in surface waters. Accordingly, the processes included in the model are advection, diffusion, point source loading of −4 Fe(CN)−4 6 and free cyanide, photolysis of Fe(CN)6 to form HCN, water–air mass transfer of free cyanide (volatilization), and biouptake/utilization/decay/biotransformation of HCN by an aquatic microflora/macrophyte community. The initial cyanide loading to the system also provides the upstream boundary condition for the system. Encoded, the computational framework developed to allow a numerical simulation of these processes was named the Cyanide Transport and Fate Model or “CTFM.” Subsequent to encoding, a site-specific calibration of the CTFM was performed, using data available from monitoring surface waters near an industrial site in eastern Tennessee [35,37].

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9.3.1 MODELING FRAMEWORK FOR NATURAL AQUATIC SYSTEMS — THE MASS BALANCE APPROACH Equation (9.14) represents one-dimensional chemical transport in a surface receiving water, in which concentration varies with time, due only to the transport processes of advection and dispersion: ∂C ∂C ∂ 2C = −u + Dx 2 ∂t ∂x ∂x

(9.14)

where C is the concentration of chemical of interest [M L−3 ], u the advection velocity of the channel [L t−1 ], Dx the longitudinal dispersion coefficient [L2 t−1 ], x the distance [L], and t the time [t]. Inclusion of a first-order reaction, in this case representing a decay process, can be added to yield Equation (9.15): ∂C ∂C ∂ 2C = −u + Dx 2 − kC ∂t ∂x ∂x

(9.15)

where k is the first-order reaction rate coefficient for the decay of the chemical of concern [t−1 ]. The reaction term may be revised to include all desired processes, such as photochemical decay of Fe(CN)−4 6 to yield HCN, loss of HCN by volatilization, and a biological uptake or decay reaction for HCN. Thus, two chemical constituents are of concern, HCN and Fe(CN)−4 6 , which are interdependent, or coupled, due to the photolysis reaction. Accordingly, when applied to the fate and transport of free cyanide (HCN) and Fe(CN)−4 6 (abbreviated for notational convenience as FeCN in the following equations), the governing equations, shown as Equations (9.16) and (9.17), obtain from Equation (9.15): ∂CFeCN ∂CFeCN ∂ 2 CFeCN = −u + Dx − kph CFeCN ∂t ∂x ∂x 2 ∂CHCN ∂ 2 CHCN ∂CHCN = −u + Dx + 6kph CFeCN − (kb + kv )CHCN ∂t ∂x ∂x 2

(9.16) (9.17)

−3 where CFeCN , CHCN are the concentrations of Fe(CN)−4 6 and HCN, respectively, mol m ; kph is −4 the first-order photolysis rate coefficient for Fe(CN)6 ; the stoichiometric coefficient of 6 implies photochemical decay of the complex goes to completion, d−1 ; kb , kv are the biodecay rate and volatilization rate coefficients for HCN, respectively. Typical boundary conditions for Equations (9.16) and (9.17), when used in a modeling application involving a riverine surface water system, could include a continuous, advective, no-dispersion upstream boundary, with a continuous or instantaneous point source of Fe(CN)−4 6 and HCN. The downstream boundary typically would display advection, and possibly, dispersion. If the latter were true, then a downstream concentration would be required, though it could be zero if appropriate to the system of interest. Application to a lacustrine surface water system would require similar conditions, although the upstream boundary may display dispersion and require a concentration condition, just as suggested above for the downstream boundary [28,40]. A widely-used approach to the application of Equations (9.16) and (9.17), whether in a riverine or lacustrine environment, involves discretization of the spatial domain to form a network of what may be visualized as completely mixed cells [28,40]. Generally, the “segmentation” scheme is based on the modeler’s judgment with respect to system uniformity, and the likelihood of spatial gradients in contaminants or other properties of interest, which may require a finer-scaled segmentation. Subsequent to discretization of the spatial domain, finite difference representations for the governing

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equations are written for each segment of the spatial grid, and a suitable solution scheme is applied to integrate numerically the resulting set of coupled equations. Applying such a discretization approach to a hypothetical spatial domain, and then selecting an arbitrary interior segment of interest, say the ith segment, it is possible to rewrite Equations (9.16) and (9.17) in the form of Equations (9.18) and (9.19) (see [28,40]): Vi

dCFeCN,i = Qi−1,i CFeCN,i−1 − Qi,i+1 CFeCN,i dt + Ei,i+1 (CFeCN,i+1 − CFeCN,i ) − Ei−1,i (CFeCN,i − CFeCN,i−1 ) − KpH,i CFeCN,i Vi ± WFeCN,i

Vi

(9.18)

dCHCN,i = Qi−1,i CHCN,i−1 − Qi,i+1 CHCN,i dt + Ei,i+1 (CHCN,i+1 − CHCN,i ) − Ei−1,i (CHCN,i − CHCN,i−1 ) + 6KpH,i CFeCN,i Vi − (Kv,i + Kb,i )CHCN,i Vi ± WHCN,i

(9.19)

where Vi is the volume of segment i, assumed to be constant with time m3 ; Qi−1,i , Qi,i+1 are the upstream flow into and downstream flow out of segment i, m3 d−1 ; CFeCN,i−1 , CFeCN,i are the concen−3 (these are assumed trations of Fe(CN)−4 6 in the center of segments i − 1 and i, respectively, mol m to represent concentration at the interface between segment i − 1 and segment i, and the interfacial concentration between segments i and segment i + 1, respectively, which constitutes a backward differencing scheme for the numerical solution to the differential equations); CHCN,i−1 , CHCN,i are the concentrations of HCN in the center of segments i − 1 and i, respectively, mol m−3 ; K pH,i the first-order photolysis rate coefficient for hexacyanoferrate in ith segment, d−1 ; K v,i the volatilization rate coefficient for hydrogen cyanide, d−1 ; K b,i the biodecay rate coefficient for hydrogen cyanide; d−1 ; WFeCN,i , WHCN,i the “loads,” mass loading rates of sources of hexacyanoferrate and hydrogen cyanide, respectively, mol d−1 ; and t the elapsed time, d. Further discussion of the finite-difference approach (Equations [9.18] and [9.19]) as a solution to the Eulerian governing equations (Equations [9.16] and [9.17]), and an example calculation for cyanide transport are available elsewhere [35]. Equations (9.18) and (9.19) represent an implementation of one of several finite-difference approaches to solving the governing mass balance equations, any of which may work well, depending on the characteristics of the system. This is particularly true for systems such as lakes, ponds, and large slow-moving rivers and streams, in which transport is not dominated by advection. In applications involving a high level of advective transport, however, the finite difference approach may require undesirable amounts of computation, to avoid errors associated with the finite difference approximations to the continuous partial differential equations that specify the mass balance problem (Equations [9.16] and [9.17]). Alternatives to the finite difference approach do exist, and can be applied to advection-dominated systems, avoiding some of the computation-related issues that can accompany the Eulerian framework. Kavanaugh et al. [35] used the Lagrangian parcel method [41–43], to simulate the transport and fate of cyanide in surface waters. This approach is explored further here. The Lagrangian parcel method (LPM) provides an alternative that involves a logical, yet conceptually quite different, approach from the Eulerian mass balance to simulate transport and fate, but it has advantages that will be noted presently. Within the conceptual framework used for the LPM, the mass of a pollutant of interest is distributed across an ensemble of small parcels, either instantaneously or at a rate equal to the mass-loading rate for a continuous source. Associated with each parcel, therefore, are a mass and a set of timedependent spatial coordinates. During each time step, all parcels are displaced according to a drift velocity (advection) component and a fluctuation velocity component at their respective locations.

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The fluctuation component of velocity is considered to be caused by turbulent diffusion, and can be represented by the characteristics of a random walk, given by:  uran =

2Dx Rnx

t

(9.20)

where uran is the random walk velocity in the direction of flow, [L t−1 ], Dx the contaminant diffusivity in the direction of flow (in one dimensional domain, Dx is the longitudinal dispersion coefficient), [L2 t−1 ], Rnx the normally distributed random number (N (0,1)), and t the time step. The fluctuation displacement for a given parcel, therefore, is the product of the time step and the fluctuation velocity. Thus, the location of the ith parcel after the jth time step, xi, j , depends on its location at the end of the previous time step, xi,j−1 , and the algebraic sum of the drift and fluctuation components of motion (u + uran,i,j ) that the parcel experienced over the time step, j t. The new location of the ith parcel, therefore, may be written as Equation (9.21): xi, j = xi, j−1 + j t (u + uran,i, j )

(9.21)

After all parcels have been displaced during a given time step, the mass of the pollutant associated with each parcel undergoes adjustment to a new value, according to the appropriate fate processes included in the model. For CTFM, the pollutants include Fe(CN)−4 6 and HCN, and the processes −4 include photolysis (Fe(CN)6 decreases, HCN increases), volatilization (HCN decreases), and a lumped biochemical reaction (HCN decreases). A linked list approach [43] provides parcel tracking and reassignment as required, to ensure mass conservation in the coupled system (i.e., hexacyanoferrate photolysis yielded free cyanide). Finally, after the masses are adjusted, the concentration of pollutant in each grid cell of the discretized spatial domain can be computed by summation of the mass of each pollutant over all parcels within each cell volume. It should be noted that the LPM compares favorably with the finite difference method for contaminant transport and fate simulation for several reasons, listed here: • The LPM is inherently stable, in contrast to some finite difference methods, because LPM time steps are open to user needs, and are simply based on the grid size and local flow velocity. • The LPM results in less numerical dispersion than the finite difference method. This is because computations by the LPM do not depend on approximations to the governing equations. • LPM simulations are more efficient than those of the finite difference method, because the computations follow pollutant parcels, and do not require the solving of a system of coupled equations. • The LPM simplifies managing spatial and temporal variation in parameters, which simplifies extending the model from one- to two- and three-dimensional problems. These advantages led to incorporation of the LPM algorithm in the CTFM for simulating cyanide transport and fate. Other details on the use of the LPM for cyanide modeling in surface waters may be found in Kavanaugh et al. [35].

9.3.2 CASE STUDY: FATE OF CYANIDE IN A SMALL STREAM ENVIRONMENT After developing and performing computational consistency checks of the CTFM as configured with the LPM approach, implementation of the model for an actual system was required. Implementation

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for a site of interest, however, required a significant amount of data about conditions at that site. Generally, the data requirements of the CTFM include information about system geometry, distances of travel, longitudinal variation in quantities and capacities of interest, including cross-sectional areas velocities, channel depth, concentrations of Fe(CN)−4 6 and HCN, pH, temperature, wind speed, solar insulation, vegetative cover, and more. Shown in Figure 9.2 is a schematic diagram of the study area, showing the main sources of cyanide in the study system (two springs), their drainage channels, and the receiving water, a small stream. Table 9.2 sets forth values for the essential physical characteristics of the system. Other information that was needed to execute the model, or to understand the approach taken for the simulation included: • The flow path from Spring #1 to the Receiving Water (see Figure 9.2) was quite short, approximately 30 m, and the residence time of water from the spring, prior to arrival in the Receiving Water, was on the order of minutes. It was assumed, therefore, that the essential character of Spring #1 was that of a point source of cyanide to the Receiving Water. • The flow path from Spring #2 to the Receiving Water (see Figure 9.2) was approximately 470 m, and water transit of this reach required on the order of hours. It was assumed, therefore, that sufficient time was available within the Spring #2 Drainage Channel for significant reaction, prior to the channel’s confluence with the Receiving Water. Hence,

Spring #2 Drainage Channel

Spring #2 (CN Source)

Pilot Treatment Wetland Cell #1

Pilot Treatment Wetland Cell #2

Receiving Water

Spring #1 Drainage Channel

Spring #1 (CN Source)

FIGURE 9.2 Schematic diagram of the study area, showing the two springs that form the main sources of cyanide in the study system, their drainage channels, and the Receiving Water.

TABLE 9.2 Average Observed Hydrologic and Meteorological Conditions for Spring #2 Drainage Channel and Receiving Water

Spring #2 drainage channel Receiving water

Temperature (◦ C)

Discharge (m3 d−1 )

Depth (m)

Velocity (km d−1 )

Wind speed (m s−1 )

13 13

1,470 23,500

0.087 0.273

16.6 17.6

2.45 2.45

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the longitudinal distribution of concentration within the Spring #2 Drainage Channel also was simulated by the CTFM, which provided a second set of simulations to check the model’s effectiveness and accuracy. • During the period of observation, the direct effect of the pilot treatment wetland system (see Figure 9.2) on the fate of cyanide in the Spring #2 Drainage Channel was not substantial. The flow diverted from Spring #2 into the wetland system, an average of 5.15 m3 d−1 , constituted less than one percent of the total flow from Spring #2. The pilot wetland system proved effective at removing cyanide from the water diverted from the Spring #2 effluent, and because it provided a dynamic treatment environment, the wetland system proved to be a valuable source of calibration information for process rates. Details of the procedures used to extract preliminary process coefficients from the wetland data for CTFM calibration are given elsewhere [35]. 9.3.2.1 Calibration Approach — Average Field Conditions The goal of the model calibration procedure was development of a numerical simulation that mimicked observed system behavior with respect to cyanide concentrations, with acceptable accuracy. The procedure consisted of iteratively executing numerical simulations, using the CTFM, with parameters and coefficients that were systematically adjusted for each iteration. The simulated cyanide concentration results for a given parameter and coefficient set were assessed for similarity, based on correlation with observed values. Monitoring program details, including the analytical methods used to acquire the data used for the CTFM calibration, are provided in Kavanaugh et al. [35]. With respect to the calibration exercise, it is important to note the following details: • Longitudinal dispersion coefficients were estimated by analysis of tracer test results. • The HCN volatilization rate was estimated from theoretical considerations, including the Henry’s law constant and site-specific conditions of temperature, wind speed, water velocity, and water depth. • The coefficient for the rate of dissolved iron cyanide photolysis (assumed to be mostly in the form of Fe(CN)−4 6 ) was estimated directly by the calibration procedure, because, photolysis essentially was independent from other fate processes. The photolysis rate coefficient estimated by this calibration procedure, however, is empirical. It is site-specific, and the potential to generalize it to other sites of interest is uncertain. Another uncertainty that underlies the photolysis rate estimate is the fact that the measured concentrations −3 of iron cyanide actually may have been a mixture of both Fe(CN)−4 6 and Fe(CN)6 , but the analytical procedures used for cyanide speciation did not permit a separate quantification of the two species (see analytical methods for complex cyanides in Chapter 7). Although there is reason to believe the source waters (the two springs) contained only −3 Fe(CN)−4 6 , or just negligible amounts of Fe(CN)6 , it is possible that some oxidation of −4 −3 Fe(CN)6 to Fe(CN)6 occurred during transport to the receiving water, because such a reaction is favored thermodynamically in the environment of interest (oxic). The importance of this is not known, but it should be noted, because the absorbance spectra and photolytic reactions between the two relatively stable iron cyanide species, Fe(CN)−4 6 and Fe(CN)−3 are known to be different [21,26]. Moreover, it is virtually certain that the 6 less stable hydrolysis species, formed by photoaquation of each component of dissolved iron cyanide, also display differing rates of reaction. It is appropriate to note, therefore, that the calibration procedure yielded a rate of photolysis of the total hexacyanoferrate −3 concentration (i.e., Fe(CN)−4 6 and Fe(CN)6 ), because of these analytically unresolved differences.

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• The rate of HCN biouptake/utilization/decay/transformation was estimated as the difference between the total HCN loss rate and the volatilization rate. The former consisted of both biochemical and water–air mass transfer processes, and was quantified by calibration ktotal = (kb + kv ), while the latter could be calculated directly (kv ). • The loadings of complex and free cyanide were estimated by calibration, because the calibrated load was considered more accurate than averaging of the observed loads for achieving representative loadings, due to a lack of information about the temporal variability of loads. The loads of Fe(CN)−4 6 and HCN estimated by this procedure, reported below, were bracketed by the range of loads observed for the system. The CTFM was developed to provide a time-variable simulation of cyanide transport and fate, which can be particularly useful when required for examining response trends to specific remediation measures and the effects of time-variable loadings or other dynamic phenomena. During the model implementations discussed here, however, all loads and other conditions were held constant with time, and the model simulation run times were selected to be sufficient to allow the predicted longitudinal concentration distribution of Fe(CN)−4 6 and HCN along the stream to reach steady state. Calibration runs were executed in two groups. One group was for the Spring #2 Drainage Channel, the results from which were used to estimate the loading to the Receiving Water from Spring #2. The second group of calibration runs was executed for the entire length of the Receiving Water, and included point source loads from both Spring #1 and #2. 9.3.2.2 Calibration Results The results of the CTFM calibration runs are summarized by the longitudinal concentration profiles presented in Figure 9.3 and Figure 9.4, and the coefficient and loading values given in Table 9.3. The CTFM results plotted in Figure 9.3 and Figure 9.4 (open symbols) show a characteristically erratic behavior that clearly reflects the effects of fluctuation velocity as a component of transport.

35 Average Conditions for Spring #2 Drainage Channel

Concentration, mg/m3

30 25 20 15 10

Observed FeCN

Modeled FeCN

Observed HCN

Modeled HCN

5 0 400

300

200

100

0

Distance from Confluence with Receiving Water (m)

FIGURE 9.3 Observed and modeled concentrations of Fe(CN)−4 6 and HCN in Spring #2 Drainage Channel (see Figure 9.2) for average conditions; vertical error bars on observed concentrations signify ±1 standard error of the mean.

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70

Average Conditions for Receiving Water

Concentration, mg/m3

60 50 Observed FeCN

40

Modeled FeCN 30 Observed HCN 20 Modeled HCN 10 0 600

500

400

300

200

100

0

–100

–200

–300

–400

–500

Distance from Confluence with Spring #2 Drainage Channel (m)

FIGURE 9.4 Observed and modeled concentrations of Fe(CN)−4 6 and HCN in Receiving Water (see Figure 9.2) for average conditions; vertical error bars on observed concentrations signify ±1 standard error of the mean.

TABLE 9.3 Calibration Values for Reaction Rate Coefficients in Drainage Channels and Receiving Water; Measured Cyanide Load from Springs #1 and #2

Spring #2 drainage channel Receiving water

Kph (d−1 ) 20 10

Kv (d−1 ) 9.3 2.7

Kb (d−1 ) 46.7 34.7

Fe(CN)−4 6 load (g d−1 ) 47.5 1430a 70.0b

HCN load (g d−1 ) 6.9 138a 27.6b

a Cyanide input to Receiving Water from Spring #1. b Cyanide input to Receiving Water from Spring #2.

The solid lines shown on each of these figures provide a smoothed, clear illustration of the predicted concentration trends. Figure 9.3 shows the final results for the calibration of the CTFM for Fe(CN)−4 6 and HCN transport and fate in the Spring #2 Drainage Channel. The observed concentration data plotted for Fe(CN)−4 6 and HCN, shown in Figure 9.3, represent the average and standard error for all observations at the indicated sampling locations along the Spring #2 Drainage Channel, as reported by Kavanaugh et al. [35]. The distance measurements shown on the abscissa of Figure 9.3 represent distances upstream (in m) from the confluence of the Spring #2 Drainage Channel with the Receiving Water. As these data indicate, the cyanide source — Spring #2 — was approximately 470 m upstream from the confluence. The results for the CTFM calibration with concentration data for Fe(CN)−4 6 and HCN obtained from the Receiving Water (Figure 9.2) are shown in Figure 9.4. Temporally averaged concentration values from the Receiving Water were used as calibration targets, similar to the approach to model calibration used for the Spring #2 Drainage Channel, and these are the values plotted as observed

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values with standard errors in Figure 9.4. It should be noted that the origin for the distance values given on the abscissa in Figure 9.4 is the confluence with the Spring #2 Drainage Channel; upstream values are shown as positive, and downstream values are shown as negative. The calibration results for the two simulated streams show reasonably good agreement with the available observed values, and the CTFM appears to reflect the magnitude of, and follows the trends in, concentration for the cyanide species of interest. The values for the CTFM parameters that were estimated by the calibration procedure show differences between the two systems that are consistent with physical and chemical differences between the systems. The calibration estimates −1 for the Receiving Water and the of the Fe(CN)−4 6 photolysis rate coefficients were 10 and 20 d Spring #2 Drainage Channel, respectively. These data indicate that the first order rate of photolysis in the Spring #2 Drainage Channel exceeded that in the Receiving Water. This is consistent with the hypothesis that the average rate of photolysis per unit of water column volume should decrease with increased water depth, due to the depth-dependent loss of radiation intensity arising from energy absorption and reflection. Indeed, the Receiving Water, on average, was much deeper than the Spring #2 Drainage Channel. The calibration also indicated that the total HCN loss rate coefficient was estimated to be 37 and 56 d−1 for the Receiving Water and the Spring #2 Drainage Channel, respectively. The volatilization rate coefficient for HCN was estimated from theory to be 2.7 and 9.3 d−1 , respectively, for the two systems. These sets of values differ by a factor of approximately 3, reflecting differences in water depth (approximately three-fold greater for the shallower Drainage Channel compared to the deeper Receiving Water) for the two systems. The calibration estimates of the overall HCN loss rates and the volatilization rates yielded firstorder rates of biouptake/utilization/decay/transformation. Rate coefficients were about 34.7 and 46.7 d−1 for the Receiving Water and the Spring #2 Drainage Channel, respectively. These biochemical reaction rate coefficients, too, are consistent with more efficient mixing and oxygen transfer for enhanced biochemical reactivity in the Drainage Channel, as compared to the Receiving Water. Nevertheless, the biotically-influenced reaction rate coefficients seem very large in comparison to the biological treatment rate coefficients commonly associated with contaminant removal in treatment systems. Snider [37], however, observed first-order biological utilization rate coefficients for HCN (in vitro, suspended culture systems) that ranged from 0.2 to 70 d−1 (average 9.4 d−1 ), for a designed set of factorial treatment combinations involving temperature and initial HCN concentration. This range spans the value determined through calibration of the CTFM. Other work has been conducted to observe biodecay rates in situ in a treatment wetland using an approach that involves microcosm isolation and monitoring of the fate of HCN and iron cyanide spikes (see Chapter 24). The loads of Fe(CN)−4 6 and HCN to each stream, estimated by CTFM calibration for the average of observed conditions, are shown in Table 9.3. For these conditions, the Fe(CN)−4 6 and HCN loads from Spring #2 to its Drainage Channel were determined to be approximately 47.6 and 6.9 g d−1 , respectively. Similarly, the loads of Fe(CN)−4 6 and HCN from Spring #1 to the Receiving Water were determined to be approximately 1430 and 138 g d−1 , respectively. In addition, the loads of Fe(CN)−4 6 and HCN from the Spring #2 Drainage Channel to the Receiving Water at the confluence were determined to be approximately 70.0 and 27.6 g d−1 , respectively. Two observations with respect to the mass loadings to the study systems are apparent from these results. First, it is clear that the cyanide loads from Spring #1 to the Receiving Water greatly exceeded the loads that reached the Receiving Water from Spring #2 via its Drainage Channel. Of greater interest, however, is the observation that the loads of Fe(CN)−4 6 and HCN to the Receiving Water from the Spring #2 Drainage Channel at the confluence were greater than the direct loads from Spring #2 to the Spring #2 Drainage Channel, by factors of 1.5 and 4.0, respectively. These differences — increases — suggest the presence of additional, unmeasured, and heretofore undetected, sources of cyanide to the drainage system, and provide a basis for further work to locate and characterize sources of cyanide contamination in the local environment.

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The results displayed in Figure 9.3 and Figure 9.4 show a reasonably good fit between the calibrated CTFM results and the observed system concentrations. Overall, the coefficients of determination (r 2 ) for the two CTFM calibrations were 0.9562 (n = 10, α < 0.001) and 0.8915 (n = 8, α < 0.001), respectively. Thus, the CTFM provided a simulation of the coupled behavior of Fe(CN)−4 6 and HCN in a surface water system that would be acceptable for application to design and management issues, and for preliminary decision support. Nevertheless, it must be recognized that significant uncertainty remains due to the small scale of the field site used for the calibration work. The small size means the system is prone to respond rapidly to meteorological events, such as rainfall and snowmelt, which makes it difficult to characterize in terms of average values. Another major source of uncertainty was the lack of information about the biouptake/utilization/decay/transformation rates in situ. More to the point, a continuing lack of information on these processes is likely to be a major ongoing concern in future work, because the comparative magnitude of the bioprocess rate coefficient indicates that its effect dominates the overall loss of cyanide from the system.

9.4 SUMMARY AND CONCLUSIONS • The principal cyanide fate-determining processes in surface waters are photolysis −3 (for Fe(CN)−4 6 and Fe(CN)6 ), volatilization (HCN), and biouptake/utilization/decay/ biotransformation (HCN). These processes are understood in terms that are sufficiently quantitative to permit numerical simulation of cyanide species fate, with a degree of accuracy that is sufficient to inform environmental management decisions. • To model cyanide fate for a system of interest, however, requires characterization of processes dynamics for the location of interest by collection of site-specific data, in the same way modeling support operations are required for modeling applications that involve other environmental contaminants. • Motivating the current study was the concern that contamination of a surface water by photolytically-active Fe(CN)−4 6 (hexacyanoferrate) would lead to significant accumulations of HCN as the photolysis reaction proceeded. This did not occur for the system of interest; rather, the combination of biologically-mediated reactions and water–air mass transfer prevented HCN from accumulating, to a large degree. • The successful development and application of the CTFM to a small-scale wetland-stream system suggests that the potential exists to use waste load allocation strategies for cyanide, as is currently done in managing the discharge of such conventional pollutants as heat, BOD, phosphorus, nitrogen, and suspended solids. • Potentially useful future modeling studies to pursue with the calibrated CTFM would focus on estimation of the loads of Fe(CN)−4 6 that would be required to be a genuine threat to the resident biota of a system, due to HCN accumulation from photolysis, and consideration of the effects of multiple point loads and distributed, or nonpoint, loads.

REFERENCES 1. Medine, A.J., Martin, J.L., and Sopher, E., Development of the speciation-based metal exposure and transformation and assessment model (META4): Application to copper and zinc problems in the Alamosa River, Colorado, in Chemicals in the Environment: Fate, Impacts and Remediation, ACS Symposium Series 806, Lipnick, R.L., Mason, R.P., Philips, M.L., Pittman, C.U., Eds., Oxford University Press, New York, 2002, p. 150. 2. UNEP, The cyanide spill at Baia Mare, Romania: before, during, and after, United Nations Environment Programme, The Regional Environmental Center for Central and Western Europe, Szentendre, Hungary, 2000. 3. Ford-Smith, M.H., The Chemistry of Complex Cyanides: A Literature Study, Department of Scientific and Industrial Research, National Chemical Laboratory, London, 1964.

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4. Doudoroff, P., Toxicity to fish of cyanides and related compounds: a review, EPA-600/3-76-038, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1976. 5. Hartung, R., Differential toxicity of forms of cyanide, Aluminum Association, Inc., Washington DC, 1990. 6. ACC, The Chemistry of the Ferrocyanides, American Cyanamid Co., New York, NY, 1953. 7. Sharpe, A.G., Chemistry of Cyano Complexes of the Transition Metals, Academic Press, New York, 1976. 8. Sillen, L.G. and Martell, A.E., Stability constants of metal-ion complexes, Special Publication Number 17, The Chemical Society, London, 1971. 9. Doudoroff, P., Some experiments on the toxicity of complex cyanides to fish, Sewage Ind. Wastes, 28, 1020, 1956. 10. Broderius, S.J., Determination of molecular hydrocyanice acid in water and studies of the chemistry and toxicity to fish of metal–cyanide complexes. Ph.D. thesis, Department of Fisheries and Wildlife, Oregon State University, Corvallis, OR, 1973. 11. Doudoroff, P. and Katz, M., Critical Review of literature on the toxicity of industrial wastes and their components to fish. I. Alkalies, acids and inorganic gases, Sewage Ind. Wastes, 22, 1432, 1950. 12. Doudoroff, P., Leduc, G., and Schneider, C.R., Acute toxicity to fish of solutions containing complex metal cyanides, in relation to concentrations of molecular hydrocyanic acid, Trans. Am. Fish. Soc., 95, 6, 1966. 13. Asperger, S., Murati, I., and Pavlovic, D., Kinetics and mechanism of the decomposition of complex cyanides of iron(II) and molybdenum(IV), J. Chem. Soc., 730, 736, 1960. 14. MacDiarmid, A.G. and Hall, N.F., Illumination pH effects in solutions of complex cyanides, J. Am. Chem. Soc., 75, 5204, 1953. 15. Moggi, L., Bolleta, F., Balzni, V., and Scandola, F., Photochemistry of coordination compounds XV: Cyanide complexes, J. Inorg. Nucl. Chem., 28, 2589, 1966. 16. Foster, G.W.A., The action of light on potassium ferrocyanide, J. Chem. Soc., 89, 912, 1906. 17. Schwartz, R. and Tede, K., Uber die photochemie der komplexverbindungen, III: die hexacyanokomplexe des dreiwertigen eisens kobatls chroms und mangans, Ber. Dtsch. Chem. Ges., 60, 69, 1927. 18. Asperger, S., Kinetics of the decomposition of potassium ferrocyanide in ultra violet light, Trans. Faraday Soc., 48, 617, 1952. 19. Mitra, R.P., Jain, D.V.S., Bannerjee, A.K., and Raghavachars, K.V., Role of free radicals in the photo oxidation of Fe+2 in acidic solutions of ferrocyanides, Nature, 200, 163, 1963. 20. Balzani, V. and Carassiti, V., Photochemistry of Coordination Compounds, Academic Press, London, 1970. 21. Kuhn, D. and Young, T.C., Direct photolysis of hexacyanoferrate(II) under conditions representative of surface waters, Chemosphere, 60, 1222, 2005. 22. Helz, G.R., Zepp, R.G., and Crosby, D.G., Aquatic and Surface Photochemistry, Lewis Publishers, Boca Raton, FL, 1994. 23. Shirom, M. and Stein, G., Excited state chemistry of the ferrocyanide ion in aqueous solution II. Photoaquation, J. Chem. Phys., 55, 3379, 1971. 24. Shirom, M. and Stein, G., Excited state chemistry of the ferrocyanide ion in aqueous solution I. Formation of the hydrated electron, J. Chem. Phys., 55, 3372, 1971. 25. Gaspar, V. and Beck, M.T., Kinetics of the photoaquation of hexacyanoferrate (II) ion, Polyhedron, 2, 387, 1983. 26. Broderius, S.J. and Smith, L.L., Direct photolysis of hexacyanoferrate complexes: Proposed applications to the aquatic environment, EPA-600/3-80-003, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1980. 27. Zepp, R.G. and Cline, D.M., Rates of direct photolysis in aquatic environment, Environ. Sci. Technol., 11, 359, 1977. 28. Chapra, S.C., Surface Water-Quality Modeling, McGraw-Hill, New York, 1997. 29. Mills, W.B., Porcella, D.B., Ungs, M.J., Gherini, K.V., Summers, K.V., Mok, L., Rupp, G.L., Bowie, G.L., and Haith, D.A., Water quality assessment: a screening procedure for toxic and conventional pollutants in surface and ground water, EPA/600/6-85/002.a, U.S. Environmental Protection Agency, Office of Research and Development, Athens, GA, 1985.

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30. Leduc, G., Pierce, R.C., and McCracken, I.R., The effects of cyanides on aquatic organisms with emphasis upon freshwater fishes, NRCC No. 19246, Associate Committee on Scientific Criteria for Environmental Quality, National Research Council of Canada, Ottawa, 1982. 31. Milne, D., Disposal of cyanides by complexation, Sewage Ind. Wastes, 22, 1192, 1950. 32. Lewis, W.K. and Whitman, W.G., Principles of gas adsorption, Indus. Eng. Chem., 16, 1215, 1924. 33. Bodek, I., Ehntholt, D.J., Glazer, A.E., Loreti, C.P., and Lyman, W., Chemical classes, in Environmental Inorganic Chemistry, Bodek, I., Lyman, W., Reehl, W.P., and Rosenblatt, D.H., Eds., Pergamon Press, New York, 1988, p. 10. 34. Yoo, K., Regression of constrained parameters from nonlinear thermodyanic models, Korean J. Chem. Eng., 3, 77, 1986. 35. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 36. Liss, P.S. and Slater, P.G., Flux of gases across the air–sea interface, Nature, 247, 181, 1974. 37. Snider, J.S., A study of the environmental fate of common inorganic cyanide species in surface waters. MS thesis, Clarkson University, Department of Civil and Environmental Engineering, Potsdam, NY, 2001. 38. Young, T.C., Issues pertaining to environmental transport, fate and biotic exposure to complex cyanides in surface waters, Technical Report, ALCOA Technical Center, Alcoa Center, PA, 1995. 39. Grady, C.P.L., Daigger, G.T., and Lim, H.C., Biological Wastewater Treatment, 2nd ed., Marcel Dekker, Inc., New York, 1999. 40. Thomann, R.V. and Mueller, J.A., Principles of Surface Water Quality Modeling and Control, Harper and Row, New York, 1987. 41. Martin, J.L. and McCutcheon, S.C., Hydrodynamics and Transport for Water Quality Modeling, Lewis Publishers, Boca Raton, FL, 1999. 42. Yapa, P.D. and Zheng, L., Simulation of oil spills from underwater accidents I: Model development, J. Hydraul. Res., 35, 673, 1997. 43. Yapa, P.D., Zheng, L., and Kobayashi, T., Linked lists for transport simulations using Lagrangian parcels, J. Comput. Civ. Eng., 10, 88, 1996.

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and Transport of 10 Fate Anthropogenic Cyanide

in Soil and Groundwater Rajat S. Ghosh, Johannes C.L. Meeussen, David A. Dzombak, and David V. Nakles

CONTENTS 10.1

Distribution and Speciation of Cyanide at Industrial Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.1.1 Cyanide in Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.1.2 Cyanide in Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.2 Impact of Site Geochemistry on Speciation and Distribution of Cyanide . . . . . . . . . . . . . . . . 10.3 Important Fate Processes for Cyanide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1 Physicochemical Fate Processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1.1 Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1.2 Adsorption. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1.3 Photodissociation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1.4 Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1.5 Phytoextraction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.2 Biological Fate Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4 Transport of Cyanide in Soil and Groundwater at Contaminated Sites . . . . . . . . . . . . . . . . . . . 10.4.1 Cyanide Mobility in Soils. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.2 Cyanide Mobility in Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

192 192 193 193 196 196 196 196 198 198 199 199 200 200 202 205 205

Cyanide has been observed as a soil and groundwater contaminant at various current and former industrial sites, including electroplating facilities, aluminum production plants, manufactured gas plants (MGP), steel plants, and metals mining and ore heap leaching facilities. It is one of the most common chemicals found at Superfund sites in the United States [1]. Cyanide has also impacted groundwater as a result of using road salt, which sometimes contains iron–cyanide anticaking agents, for deicing purposes during winter months [2]. At many industrial sites with cyanide-bearing wastes, including aluminum manufacturing and MGP facilities, the leaching of iron–cyanide solids, which are present as onsite fill, can yield detectable concentrations of dissolved cyanide in groundwater. Because of the nature of the cyanide source, dissolved cyanide in groundwater at aluminum facilities and MGP sites consists primarily of soluble iron–cyanide complexes [3–5]. However, in other instances, such as at ore heap leaching sites or electroplating facilities, weak-acid-dissociable (WAD) cyanide complexes and free cyanide are also present, in addition to the iron–cyanide complexes, depending on site conditions. 191

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The fate and transport of cyanide compounds in soil and groundwater have been studied by researchers in Europe and the United States [3,5–11]. From this work, it has been found that the soluble strong metal–cyanide complexes, like iron and cobalt cyanides, are stable in the dark at neutral to high pH [4], and are highly resistant to biodegradation [12,13]. Free cyanide, on the contrary, is an excellent source of nitrogen for fungal and bacterial growth, and is readily amenable to microbial degradation ([14–18] and Chapter 6). In addition, at mildly acidic to neutral pH conditions, many metal–cyanide complexes, including iron–cyanide species, exhibit some degree of adsorption onto both natural organic and inorganic adsorbents (see Chapter 5). Combined, these observations suggest that dissolved cyanide may or may not be mobile and persistent in groundwater systems, depending on site conditions. The fate and transport of cyanide compounds in soil and groundwater must be understood to conduct risk assessments, and to design control and remediation measures for cyanide at impacted sites. This chapter highlights and discusses the principal fate processes of dissolved cyanide in groundwater, based on a combination of literature data and site-specific field studies. The influence of these processes on the transport of cyanide in the subsurface environment is also examined.

10.1 DISTRIBUTION AND SPECIATION OF CYANIDE AT INDUSTRIAL SITES Cyanide is often present in industrial byproducts or residuals as solid phase iron–cyanide compounds (Chapters 4 and 26). Leaching of soluble metal–cyanide complexes (primarily iron cyanides) from these materials can eventually result in cyanide impacts to groundwater. At some sites, such as electroplating and ore heap leaching sites, other metal–cyanide complexes, including, copper, gold and nickel, may also be present. The specific distribution and speciation of cyanide at an industrial site is a function of the characteristics of the existing or former production processes, as well as past and present environmental conditions at the site.

10.1.1 CYANIDE IN SOIL Cyanide is used in large quantities in a number of industrial processes, and is also produced incidentally in many processes. Thus, impacts in soil have been observed at many industrial sites. At many of the sites, the predominant forms of cyanide compounds are iron–cyanide solids, the most prominent among them being Prussian Blue or ferric-ferrocyanide [Fe4 (Fe(CN)6 )3 (s)]. On the contrary, at ore heap leach and electroplating spill sites, the cyanide in soil is usually dominated by a mixture of iron–cyanide and other metal–cyanide compounds [19,20]. The source of cyanide compounds in soil and groundwater at MGP or coke production sites is usually oxide-box residuals that were managed onsite as fill [10,21]. During active plant operations, boxes containing wood shavings or crushed blast-furnace slag, mixed with a chemically active form of hydrated iron oxide, were used to remove H2 S(g) from the manufactured gas or coke oven gas. During this process, hydrogen cyanide (HCN) was removed simultaneously from the gas [22]. Over time, most of the iron oxide reacted with the H2 S(g), and was removed for regeneration prior to reuse. This regeneration/reuse cycle was repeated until the oxide lost its effectiveness. At this point, it was replaced with fresh material. The material that was replaced was commonly called “spent oxide,” which was used as fill material, both on and offsite. The “spent oxide” material contained double-iron–cyanide compounds, like Prussian Blue, which formed over time as the free cyanide reacted with the iron. The presence of spent oxide in soil is readily apparent from the intense blue color of the Prussian Blue. The double-iron–cyanide salts, like Prussian Blue [Fe4 (Fe(CN)6 )3 (s)] and Turnbull’s Blue [Fe3 (Fe(CN)6 )2 (s)], have very low solubility under acidic to neutral pH conditions ([23,24] and Chapter 5). As a result, these compounds dominate cyanide-impacted soils that are acidic to neutral in nature. Other iron–cyanide compounds, like manganese iron (II) cyanide [Mn2 (Fe(CN)6 (s)] [25] or potassium zinc iron (II) cyanide [K2 Zn3 (Fe(CN)6 )2 (s)] [26], could be present in more alkaline

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soils. However, except for Prussian Blue, most iron–cyanide solids are hypothesized precipitates based on fitting of solubility data, and there are no independent solubility product determinations. The mixed-metal–cyanide compounds may be less soluble than double-iron–cyanide compounds at higher pH levels [11]. Dissolution of these metal–cyanide compounds under site-specific geochemical conditions will eventually release soluble iron–cyanide complexes into the groundwater. In addition to the cyanide source materials, dissolved metal–cyanide complexes can also adsorb onto various natural soil adsorbents (Chapter 5). Iron–cyanide complexes exhibit limited adsorption onto iron oxides [27,28] and sandy siliceous soil [3] in the neutral pH range, but absorb significantly on aluminum oxides [29] and kaolin clay under neutral to mildly acidic pH conditions [6]. Natural organic matter can also act as an important adsorbent for both ferro- and ferricyanide complexes over a range of pH conditions [30]. This information suggests that, in addition to dissolution of source materials such as spent oxide, desorption of iron–cyanide complexes should be considered for certain soil types.

10.1.2 CYANIDE IN GROUNDWATER Dissolved cyanide is present in groundwater as a result of the dissolution of cyanide source materials and the desorption of adsorbed metal–cyanide complexes from soil. The dissolved-phase cyanide species can exist in a number of different chemical forms, ranging from free cyanide (HCN or CN− ) to weak-acid-dissociable complexes (e.g., cyanide complexes with copper, zinc, and nickel) to available cyanides (e.g., weak-acid-dissociable complexes plus mercury–cyanide complex) to strong-acid dissociable complexes (complexes of cyanide with iron and cobalt). The exact distribution of dissolved cyanide species in groundwater varies considerably depending on the type of the cyanide source material and the site-specific geochemistry. For example, at aluminum smelting sites, salt storage facilities, and MGP sites, iron–cyanide complexes (ferro- and ferricyanide) dominate the speciation [3,7]. This is because the primary source of dissolved phase constituents is the iron– cyanide solid Prussian Blue. Only 10% or less of the cyanide at such sites exists as weak metal– cyanide complexes, of which free cyanide constitutes only a few percent [2,3,10,31,32]. On the contrary, leakages from tailing ponds associated with gold mining and heap leaching operations, as well as spills from electroplating operations, can release to soil and groundwater free cyanide and various metal–cyanide complexes formed with gold, zinc, cadmium, silver, copper, nickel, iron, and cobalt [33]. Table 10.1 lists some common cyanide species that can be present in groundwater at industrial sites, as well as an approximate assessment of their relative abundance. The properties and reactivities of these cyanide species are discussed in Chapter 5. The chemical classifications of the various dissolved cyanide complexes are addressed in detail in Chapter 7. Among the strongest metal–cyanide complexes, and often the most important in groundwater, are the iron–cyanide complexes: ferrocyanide [Fe(CN)6 ]4− , in which iron is in the +II oxidation state, and ferricyanide [Fe(CN)6 ]3− , with iron in the +III oxidation state. The ferrocyanide complex usually dominates industrial groundwater, due to reducing conditions that usually exist. A third iron–cyanide complex has also been reported [10,34] in groundwater and soil leachates at multiple MGP sites in the United States. This complex has been identified tentatively as an iron–pentacyano complex with a chemical formula of [Fe(CN)5 NHCH3 ]4− [35].

10.2 IMPACT OF SITE GEOCHEMISTRY ON SPECIATION AND DISTRIBUTION OF CYANIDE The dissolution of cyanide-containing solid residuals, desorption of cyanide complexes from soils, or spilling of aqueous solutions containing cyanide result in the release of soluble cyanide species to the groundwater. While the dissolution of almost all cyanide-containing solids, including the iron– cyanide solid Prussian Blue, is a function of pH and redox potential [5,9,23,24,36], the desorption of metal–cyanide complexes depends primarily on pH and the soil type [6,27–30].

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TABLE 10.1 Common Cyanide Species in Groundwater at Industrial Sites Industrial site type(a)

Classification

Species

MGP

Aluminum smelting

Mining/heap leaching

Electroplating

Free cyanide

HCN, CN−

Trace (<3%)

Trace (<3%)

Moderate (10–30%)

Moderate (20–40%)

Weak metal–cyanide

Cu(CN)2− 4

Low (<10%)

Low (<5%)

Moderate (20–40%)

Moderate (20–40%)

Au(CN)− 2

High (>90%)

High (>90%)

High (30–60%)

High (30–60%)

Fe(CN)4− 6 Fe(CN)3− 6 Co(CN)3− 6 SCN−

Low (<10%)

None

None

None

Zn(CN)2− 4

complexes

Ag(CN)− 2

Ni(CN)2− 4 Cd(CN)2− 4

Strong metal–cyanide complexes

Other

(a) Cyanide species distribution categories: Trace: <3%, Low: 3 to <10%, Moderate: 10 to <40%, High: 30 to <90%.

The joint effects of pH and pE (pE = 16.9 × Eh , where Eh is the redox potential in volts) control the equilibrium dissolution behavior of Prussian Blue, a cyanide-bearing solid common at cyanide-contaminated sites. In addition, there is a continuum of different forms of iron–cyanide solids that exist as the pH and pE values shift in a geochemical environment ([23] and Chapters 2 and 5). Figure 10.1 shows the predominance diagram of different iron–cyanide solids as a function of the pH and pE, groundwater in a system with hydrous ferric oxide present. As shown in this figure, pure Prussian Blue can only exist under acidic pH and oxic conditions. As pH increases, Prussian Blue coexists with hydrous ferric oxide in the form of a solid solution (coprecipitant). At high pH, Prussian Blue dissolves. Turnbull’s Blue, another iron–cyanide solid, is found to exist under anoxic conditions for a wide range of pHs. The transition from blue to green color that is often observed occurs under neutral to alkaline pH conditions, when Turnbull’s Blue forms a solid solution with hydrous ferric oxide. Results of several spectroscopic studies have suggested that Prussian Blue and Turnbull’s Blue are identical [37,38]. However, it has been found that invoking a reaction for dissolution of [Fe3 (Fe(CN)6 )2 (s)] provides the best fit for iron–cyanide equilibrium solubility data for mildly acidic to neutral pH regimes when the ORP (oxidation-reduction potential) values are low (anoxic conditions prevail) [23], while a reaction for dissolution of Prussian Blue dominated for oxic conditions (Figure 10.1). Mössbauer spectroscopic analysis also suggested a different Fe(II)/Fe(III) ratio compared to Prussian Blue for the anoxic regions, consistent with the [Fe3 (Fe(CN)6 )2 (s)] chemical formulation [23]. The redox potential also controls the dissolved iron–cyanide complex speciation in groundwater (Figure 10.2). As shown in Figure 10.2, under moderately oxic to highly oxic conditions, the ferricyanide [(Fe(CN)6 )3− ] complex dominates, while the ferrocyanide [(Fe(CN)6 )4− ] dominates under anoxic conditions. Adsorption studies conducted with cyanide complexes of Cd, Cu, Fe, Ni, and Zn on a natural soil adsorbent, goethite (α-FeOOH), indicated that for highly charged complexes, like [Fe(CN)6 ]4− and [Ni(CN)4 ]2− , adsorption is significant under acidic pH conditions (pH < 6) [28]. Similar

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Oxic conditions

25 20

Boundary of natural conditions Prussian Blue Fe4(Fe(CN)6)3

15

Solid solution of Prussian Blue and Fe(OH)3

10

Anoxic conditions

pE

Fe(OH)3 [Cyanide dissolved as iron– cyanide complexes]

5 Solid solution of Turnbull’s Blue and Fe(OH)3

Turnbull's Blue Fe3(Fe(CN)6)2

0

Solid solution of Turnbull's Blue and Fe(OH)3

–5 Boundary of natural conditions

Fe(OH)2

–10 0

2

4

6

8

10

12

pH

FIGURE 10.1 Predominance diagram of iron–cyanide solids, in a system with hydrous ferric oxide present, as a function of pH and pE. 20.0 Water oxidized (region of instability)

15.0 Fe(CN)63 –

pE

10.0

5.0

0.0 Fe(CN)64–

–5.0

Water reduced (region of instability)

–10.0 0.0 1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

10.0

11.0

12.0

pH

FIGURE 10.2 Predominance diagram of dissolved iron–cyanide complexes.

behavior was also observed in column studies performed with aluminum oxides and kaolin clay type materials under acidic pH conditions [6,29]. This is significant, considering the fact that the ferrocyanide [Fe(CN)6 ]4− complex is one of the most dominant forms of dissolved cyanide species in groundwater at industrial sites. Transport studies in fixed bed columns with sandy soil and ferrocyanide influent, however, reveal absence of any retardation, indicating no significant adsorption capacity of SiO2 (s) for ferrocyanide [3]. Also, the presence of competing adsorbates like sulfate ion can suppress the iron–cyanide complex adsorption processes with natural adsorbents like goethite, under neutral to acidic pH conditions [27]. Overall, available information on iron–cyanide complex adsorption suggests that both pH and type of soil will influence the distribution of dissolved iron–cyanide complexes between groundwater and soil. Further discussion on adsorptive properties of various cyanide species is provided in Section 10.3.1.2.

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10.3 IMPORTANT FATE PROCESSES FOR CYANIDE The fate processes affecting cyanide distribution and speciation at industrial sites fall under two broad categories: (i) physicochemical fate processes, which include the various physical and chemical mechanisms that determine the distribution of solid and dissolved forms of cyanide, and (ii) biological fate processes, which primarily include the various microbially-mediated cyanide degradation mechanisms.

10.3.1 PHYSICOCHEMICAL FATE PROCESSES The various physicochemical fate processes of interest include precipitation–dissolution of iron– cyanide solids; adsorption–desorption processes involving metal–cyanide complexes and free cyanide ions; photodissociation of iron–cyanide complexes; volatilization of gaseous HCN; and phytoextraction of dissolved metal–cyanide complexes and free cyanide ions.

10.3.1.1 Precipitation The iron–cyanide solids at industrial sites belong to the class of complex coordination compounds that are crystalline in nature, where multiple iron-cyanide complexes (ferro- or ferricyanide complexes) are bonded to a central metal cation (iron or other heavy metals like Co) (see Chapter 5). These crystalline compounds have well-defined solubility behavior. There is basic understanding of the complex interaction of pH and redox potential (pE) with the precipitation–dissolution patterns of Prussian Blue [5,9,23,24,35]. The first comprehensive set of experiments performed to study the precipitation and dissolution behavior of Prussian Blue was done by Meeussen et al. [24]. Solutions containing equimolar concentrations of chelated iron and [Fe(CN)6 ]4− complexes were used to precipitate Prussian Blue under a range of pH and pE conditions. According to Meeussen et al. [24], Prussian Blue is stable at pH < 6, and tends to dissolve readily under alkaline conditions. Ghosh et al. [23] performed a follow-up study to investigate the effect of pH and pE on the solubility behavior of Prussian Blue under equilibrium conditions. Results from this study indicated that Prussian Blue is soluble under alkaline pHs (pH > 7) and oxic (pE > 5) conditions. Figure 10.3 provides contours of dissolved total cyanide concentrations for equilibrium with Prussian Blue in a system with hydrous ferric oxide under various pH and pE conditions, based on laboratory studies conducted by Ghosh et al. [23]. As shown in this figure, the solubility of Prussian Blue is very low at low to neutral pH conditions and moderate redox levels. This explains why soils containing Prussian Blue can persist for decades in an acidic to neutral soil environment at industrial sites. However, even under alkaline pH conditions, researchers have observed that iron– cyanide compounds were present in the soil solid phase [9]. While equilibrium calculations indicate that Prussian Blue will not be stable under such conditions, the observed stability can be explained by the very slow dissolution rate of iron–cyanide solids in the dark [9,24]. Reducing conditions also have a role. For example, as shown in Figure 10.3, even under alkaline pH conditions (7 < pH < 9), Prussian Blue would only be partially dissolved, if the anoxic conditions prevail. Another possibility is that under alkaline conditions, adsorption becomes more important.

10.3.1.2 Adsorption Adsorption of metal–cyanide complexes and free cyanide ion on natural soil adsorbents has been examined in a number of studies (Chapter 5). In general, soil properties, such as low pH (pH < 5), and the presence of iron and aluminum oxides, and clay material with high anion exchange

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Oxic conditions

Boundary of natural conditions

20 15 10 pE

Anoxic conditions

197

2.98 0.57 0.11

Cyanide fully dissolved

5 15.7 0

Low dissolved cyanide

–5 Cyanide fully dissolved Boundary of natural conditions –10 0 2 4 6

15.7

8

10

12

pH

FIGURE 10.3 Total dissolved cyanide (mg/l) in equilibrium with solid solution of Prussian/Turnbull Blue and Hydrous Ferric Oxide as a function of pH and pE. (Source: Ghosh, R.S. et al., Environ. Eng. Sci., 16, 293, 1999. With permission.)

capacity, tend to increase adsorption of anionic metal–cyanide complexes in soil and groundwater environment [6,27,28,30]. Free cyanide adsorbs weakly, or not at all, to oxide minerals [28,39]. As discussed in Chapter 5, however, free cyanide does adsorb significantly on organic carbon. Higgins and Dzombak [39] showed that under near-neutral pH conditions, free cyanide adsorption on soils is correlated with organic carbon content. An empirical relationship developed from experimental data is presented in Chapter 5. Alesii and Fuller [6] conducted column experiments with five different types of soils (sand, silty till, till, silty clay, and clay) and three different influent solutions of cyanide: KCN in de-ionized water, K3 Fe(CN)6 in de-ionized water, and KCN in natural landfill leachate. The results of this study showed that the mobility of the cyanide in these studies, HCN and Fe(CN)3− 6 , was dependent on the dominant types of soil minerals and pH. For all the three forms of cyanide, soils with high positive charges (i.e., kaolin, chlorite, and gibbsite type clay), high iron oxide content, and acidic pHs seemed to be most effective in retarding the cyanide transport. On the contrary, calcareous soils with high pHs and low clay content seemed to increase the mobility of cyanide. The soils were not sterilized, so biological degradation may have influenced the fate of free cyanide in the systems. Theis and West [28] conducted adsorption studies with various metal-cyanide complexes on goethite type material, and found substantial adsorption of both ferrocyanide and nickel tetracyano complexes in the acidic pH range. Similar adsorption studies on goethite were conducted by Rennert and Mansfeldt [27], who investigated the adsorption of iron–cyanide complexes in the presence of competing sorbates like sulfate anion. According to Rennert and Mansfeldt [27], under acidic pH conditions, ferricyanide appeared to form outer-sphere, and weak inner-sphere surface complexes with goethite, as opposed to ferrocyanide, which formed stronger inner-sphere surface complexes, resulting in significantly less mobility, when compared to the ferricyanide complex. Also, the adsorption of the ferricyanide complex was highly suppressed compared to the adsorption of the ferrocyanide ion, due to the presence of competing sulfate anion, under acidic pH conditions (pH = 3.5). The effect of chloride and phosphate on the desorption of these complexes was also studied by the same authors. Only the ferricyanide complex was found to desorb because of the increase in the ionic strength of the solution, caused by the presence of high chloride concentration. However, the presence of an unbuffered phosphate solution seemed to desorb both of the iron-cyanide complexes from goethite.

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Adsorption of both forms of iron–cyanide complexes in 17 uncontaminated soils was studied by Rennert and Mansfeldt [30]. The results of this study showed that increasing soil organic matter (SOM) content of the soil (>10 g/kg) resulted in higher adsorption of both the complexes, possibly because of a reaction between the cyaonogenic nitrogen and reactive functional groups of SOM such as, for example, the quinone groups in the humic acids. For soils with less organic matter content (<10 g/kg), pH and clay content govern the adsorption behavior of these complexes. Bushey and Dzombak [29] performed ferrocyanide (Fe(CN)4− 6 ) adsorption studies on γ-alumina (γ-Al2 O3(s) ) and gibbsite (Al(OH)3(s) ) over a wide pH range and various solid loadings. The results of this study showed increasing ferrocyanide adsorption with decreasing pH, consistent with the general pH dependence for adsorption of anions on oxide minerals. Greater surface reactivity of the γ-Al2 O3(s) resulted in an almost 300-fold increase in ferrocyanide adsorption on the alumina surface, when compared to the adsorption on Al(OH)3(s) . Ferrocyanide adsorption on γ-Al2 O3(s) was significantly greater than previously reported data for goethite. The investigation showed that ferrocyanide could adsorb significantly on aluminum oxide surfaces in the acidic to alkaline pH range, with the extent of adsorption highly dependent on pH, the solid crystalline structure, and associated surface reactivity. 10.3.1.3 Photodissociation Some metal–cyanide complexes can be decomposed by photodissociation, which is relevant to groundwater only insofar as groundwater has potential for discharge to surface water (see Chapter 9). Photodissociation of metal–cyanide complexes, particularly the ferro- and ferricyanide complexes, has been reported by numerous researchers during the last five decades [24,40–47]. The photodissociation equations for ferrocyanide and ferricyanide complexes are provided in Chapter 5 and Chapter 9 of this book. Available data indicate that photodissociation of dissolved iron–cyanide complexes can be rapid. For example, Johnson et al. [45] studied the photodissociation of ferrocyanide complexes in drainage ditches associated with precious metal mining facilities, and found rapid photodissociation of ferrocyanide complexes in daylight, followed by slow or no dissociation in the absence of light. Johnson et al. showed that the decrease in iron–cyanide complexes in the open ditches was followed by an increase in WAD cyanide. The authors concluded that any photochemically produced cyanide will likely be complexed with other metals in the environment. Finally, there are various other factors that influence the photodecomposition of iron–cyanide complexes upon discharge to surface waters, including pH, sunlight intensity, temperature, turbidity, and depth of the water column. These factors are considered in detail in Chapter 9. 10.3.1.4 Volatilization Although none of the soluble metal–cyanide complexes at industrial sites have significant vapor pressures at environmental conditions, HCN is volatile, meaning that it may escape from shallow groundwater systems and move into the unsaturated zone, or may volatilize upon discharge to surface waters [48,49]. Data regarding the Henry’s Law constant for HCN are surprisingly sparse. Table 5.1 lists the commonly reported values of Henry’s Law constant for HCN. The Henry’s Law constant is a function of temperature: higher temperatures result in larger constants, which yield higher volatilization rates [48]. It is important to note that an estimate of the equilibrium distribution of HCN, as governed by the Henry’s Law constant, provides little information about the rate at which this equilibrium will be approached in a natural aquatic system. Such rate data can only be acquired through studies that are focused on the quantification of liquid-to-gas mass transfer coefficients (Chapter 9). These mass transfer coefficients could then be used to predict volatilization fluxes. Researchers have employed in-field flux chamber devices for direct measurement of HCN volatilization rates from a wetland fed by a cyanide-containing groundwater source [50]. A schematic diagram of the flux chamber,

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Effluent air line

Influent air line

Flux chamber

Flotation collar

NaOH solution Collection vessel

Vacuum pump

Wetland water surface

FIGURE 10.4 Flux chamber for measuring HCN volatilization rate from groundwater-fed wetland. (Source: Kavanaugh, M.C. et al., Water Environment Research Foundation, 2003. With permission.)

including the gas collection system, is shown in Figure 10.4. The mass flux of free cyanide using such a device is calculated by mass balance, using elapsed time of the test, the airflow rate, the measured HCN concentration, and volume of the collection solution.

10.3.1.5 Phytoextraction Research involving studies on willow hydroponics has shown that dissolved free cyanide and ferrocyanide can be taken up by certain species of plants (e.g., Willow sp.), and assimilated into plant amino acids (Chapter 24). Although the rate of assimilation is much higher for free cyanide, the research has shown that even strong metal–cyanide complexes like ferrocyanide can be taken up and metabolized in plant tissue. Based on the results of these hydroponics studies, models have been developed to design and interpret the performance of field scale phytoextraction systems [51]. While studies have been limited, available data make it clear that various plants have the ability to take up and metabolize cyanide species from the ground [52]. This is not surprising, considering the role of plants in the natural cycle of cyanide (Chapter 12). Cyanide uptake by plants may influence cyanide distribution, fate, and transport at vegetated sites having cyanide contamination above the water table, which is generally the limit of root penetration.

10.3.2 BIOLOGICAL FATE PROCESSES Biological fate processes influencing cyanide occurrence and fate in groundwater systems can be classified as (i) biological synthesis of cyanide or biological cyanogenesis, and (ii) biological degradation of natural as well as anthropogenic cyanides. At most contaminated sites, the biological degradation processes play a greater role in influencing the overall fate of cyanide, as compared to biological cyanogenesis. Hence, this section is primarily devoted to the discussion of the various biological degradation processes that utilize cyanide in a groundwater and soil environment. More in-depth discussions of biological synthesis and degradation of cyanide are provided in Chapter 3 and Chapter 6. The biological degradation of free cyanide, thiocyanate, and iron–cyanide complexes has been the subject of many investigations (Chapter 6 and Chapter 23). The majority of these investigations have been conducted under aerobic conditions. The search for organisms that can beneficially use cyanide as a growth substrate began in the area of municipal wastewater treatment. Pettet and Mills [53] described an acclimated trickling filter process that could remove 100% of free cyanide from an influent containing 100 mg/l HCN. Since then, biological treatment methods for cyanide biooxidation have been developed and piloted, using both fixed film and suspended culture systems. In

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TABLE 10.2 Common Aerobic Cyanide Degradation Reactions in Pure and Mixed Cultures Reactants KCN, H2 O, O2 HCN, H2 O HCN, H2 O HCN, H2 O HCN HCN CN− , S2 O2− 3 SCN

Intermediates

HCNOH2 , H2 O

HOCN

Products KOH, 2NH3 , 2CO2 HCOOH, NH3 , CO2 HCNOH2 , H2 O HCOOH, NH3 CO2 , NH3 CO2 , NH3 SCN− , SO2− 3 + SO2− 4 , NH4 , CO2

Enzyme Cyanide hydratase Cyanide hydratase Formamide hydrolyase Not identified Various extracellular enzymes Dioxygenase Rhodanase Unknown

Source: Kjeldsen, P., Water, Air, Soil Pollut., 115, 279, 1999.

these systems, cyanide predominantly serves as a nitrogen source for the microorganisms. Table 10.2 presents some important free cyanide biodegradation pathways established in laboratory microbial studies, using either isolated bacterial strains or mixed cultures [36]. As shown in this table, the degradation reactions are all triggered by extra-cellular enzymes, some producing intermediate organic by-products. Under most environmental conditions, free cyanide degrades to CO2 and the ammonium ion. − In the presence of thiosulfate, S2 O2− 3 , cyanide can also be transformed to thiocyanate (SCN ), which can subsequently be mineralized to CO2 , ammonium ion, and sulfate. Although metal–cyanide complexes are resistant to microbial degradation [13,54,55], there have been reports of microbial degradation of metal–cyanide complexes. For example, pseudomonas strains, isolated from a full-scale waste treatment plant associated with a gold mining operation, were reported capable of degrading cyanide, thiocyanate, and iron–cyanide complexes [56]. Mudder and Whitlock [56] reported 95% removal of total cyanide (which included iron-cyanide complexes) and 98% removal of weak-acid-dissociable cyanide in an activated sludge process. The same results were observed in a degradation experiment involving contaminated soil taken from other gold mine tailing plants [57,58]. However, the influence of the photodecomposition of iron–cyanide complexes to simpler cyanide species on these results was not investigated. It may be possible that the observed biodegradation was the result of a two-step sequence of reactions, the first of which was photodecomposition of the iron–cyanide complexes to simpler, more biologically available, cyanide species. Hommelgaard [59] demonstrated that iron–cyanide complexes leached out of MGP site residuals can be naturally degraded under aerobic conditions to CO2 and nitrate. Again, however, in the absence of photodissociation controls, it is not clear if direct biodegradation of hexacyanoferrate actually occurred. Available information from experiments with controls suggests it is likely that hexacyanoferrate first has to dissociate to free cyanide, prior to any biodecomposition.

10.4 TRANSPORT OF CYANIDE IN SOIL AND GROUNDWATER AT CONTAMINATED SITES 10.4.1 CYANIDE MOBILITY IN SOILS The mobility of soluble cyanide species in soil is governed by precipitation–dissolution, as well as adsorption–desorption reactions with the soil matrix. Biodegradation reactions and changes in solution chemistry will also influence cyanide speciation and hence transport. For example, the

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transport of free cyanide in soil solution can be influenced by surface adsorption processes, especially with high organic carbon content soils, and by reactions with metals in solution that could lead to additional kinds of adsorption reactions, or even precipitation reactions. The mobility of dissolved metal–cyanide complexes, such as soluble ferro- and ferricyanide, is controlled primarily by precipitation–dissolution and adsorption–desorption processes, as these species are less subject to biodegradation. As mentioned earlier in this chapter, adsorption can influence the movement of soluble iron–cyanide complexes in soils, depending on the mineral surfaces present. Many researchers highlight the potential importance of iron–cyanide solid colloidal transport in acidic soils [8,27]. Vertical migration of Prussian Blue solids in acidic unsaturated soils has been attributed to Prussian Blue colloidal transport by Mansfeldt et al. [8]. However, this is only a hypothesis, since the phenomenon of Prussian Blue colloidal transport in acidic soils has never been directly demonstrated. Furthermore, high ionic strength of the soil solution in the macropores can suppress such colloid movement [60]. Another possible explanation of the observed vertical distribution of iron–cyanide solids in unsaturated soils, lies in the dissolution–precipitation characteristics of Prussian Blue or any other solid phase compounds. It is known that the dissolution of cyanide from these compounds is a strong function of pH and pE of the soil solution [4,5,23]. This relationship could explain the vertical migration of cyanide in acidic subsoil layers by considering the dissolution of high amounts of ferrocyanide under low pH and high pE conditions in the upper subsurface zones, followed by vertical migration, and then re-precipitation of some of the iron–cyanide complexes when relatively low pE regions are encountered. A case study at a Wisconsin MGP site in the United States serves to illustrate how pH and redox potential can control the movement of cyanide in unsaturated soils [3,61]. At this particular site, the dissolved cyanide plume, dominated by iron-cyanide complexes, was still present after six years, following the removal of all cyanide-impacted residuals from the source area [3]. The water table at this site ranges anywhere from 12 to 28 ft below the ground surface. Soil boring analyses near the source area indicated isolated stringers of blue colored soil along the vertical section of the boring, primarily in the unsaturated zone (2 to 28 ft bgs), and in a few areas in the upper portions of the saturated zone (28 to 35 ft bgs), with total cyanide concentration in these stringers ranging from 1 to 95 mg/kg. Further correlation of field measured pH and redox potential with Prussian Blue equilibrium solubility information [23] indicated that the continuing source of cyanide to groundwater is a result of the leaching of iron–cyanide complexes. These complexes leached from the original spent oxide residuals, and then re-precipitated on sand and gravel surfaces below the original source areas under excess iron conditions. Isolated pockets of precipitated solids at different depths along the entire vertical thickness of the unsaturated zone indicated that this leaching and re-precipitation phenomenon may be occurring over the entire unsaturated zone, and may be providing the source for cyanide impacts to groundwater at this site. Similar re-precipitation events were also observed by Meeussen et al. [9] in soil profiles near the capillary fringe zone above the water table, primarily due to the change in the pH and redox potential. Ohno [62] reported an observed ferrocyanide concentration in an iron-rich soil in the vicinity of a salt storage facility that was lower than the calculated maximum, and attributed this observation to the re-precipitation of ferrocyanide complexes. Meeussen et al. [5] investigated the mobility of the iron-cyanide complexes, ferro- and ferricyanide, at 12 MGP site soils, six with high pH clay type soil (5.5 < pH < 7.2), and six with low pH sandy soils (3.5 < pH < 5.5). Figure 10.5 shows the cyanide concentrations measured in groundwater at these sites. As shown in Figure 10.5, the highest cyanide concentrations were observed in the soils with high pH levels, while the lowest concentrations of dissolved cyanide occurred in the acidic soils. This observation further validates the importance of pH in controlling the solubility of Prussian Blue and dictating its mobility in soils. Earlier modeling studies performed by Meeussen et al. [9] with field data also indicated that pH and redox potential control the mobility of cyanide in acidic soils. In alkaline soils, Prussian Blue is

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Clay soils (high pH)

35 Total CN concentration (mg/l)

Clay soils (low pH) 30 25 20 15 10 5 0

1

2

3

4

5

6

1

2

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4

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6

Site

FIGURE 10.5 Concentration of total cyanide in groundwater at 12 MGP site soils from The Netherlands. (Source: Meeussen, J.L. et al., J. Environ. Qual., 23, 785, 1994. With permission.)

soluble, and dissolved hexacyanoferrates are mobile once leaching occurs. However, the authors of this study indicated that the slow dissolution kinetics of Prussian Blue might be responsible for its persistence even in alkaline soils. In addition to precipitation–dissolution processes, adsorption–desorption of free cyanide and iron-complexed cyanides can also influence the mobility of the species as a function of pH and soil composition. Acidic soils with high concentrations of iron or aluminum oxides and clay content will adsorb metal–cyanide complexes [6,27–29]. Neutralization of acid generation capacity in these types of soils, by addition of lime or other amendments, could mobilize the cyanide species. Free cyanide can be retained by soils with appreciable organic carbon content ([39] and Chapter 5). Increasing the amount of natural organic material in soil, for example, humic and fulvic acids, could also enhance iron-cyanide complex adsorption over a wide pH range, and retard its movement [30].

10.4.2 CYANIDE MOBILITY IN GROUNDWATER The first major transport study involving free cyanide and the ferricyanide complex in groundwater was performed by Alessi and Fuller [6]. These researchers studied the mobility of free cyanide and ferricyanide complexes in laboratory fixed-bed columns packed with five different types of soils of varying physical and chemical properties. Free cyanide and ferricyanide ions in solution were found to be mobile in some soils, but were retained in some soils with low pH and the presence of free iron oxide and clay material with high positive charges. In contrast, soils with neutral to alkaline pH conditions, calcareous in nature, and with low clay content, seemed to increase the mobility of both forms of cyanide. In addition to the soil pH, clay content, iron and aluminum oxide content, and anion exchange capacities, the presence of high concentrations of soil organic matter (SOM) can also influence the mobility of cyanide in groundwater. Studies by Rennert and Mansfeldt [27,63] indicated that the presence of a high concentration of SOM could attribute to significant retardation of iron cyanide complexes in the groundwater, possibly due to surface reactions between iron–cyanide N and reactive groups of SOM.

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Total effluent cyanide conc., ppm

7 6 Influent total cyanide concentration = 6.2 ppm Site groundwater (MW 423A, July 1996) q = 0.1 ml/min Breakthrough like a nonreactive tracer

5 4 3 2 1 0 0

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Pore volumes passed through column

FIGURE 10.6 Total cyanide breakthrough in column study with MGP site groundwater and aquifer material. (Source: Ghosh, R.S. et al., Water Environ. Res., 71(6), 1205, 1999. With permission.)

High mobility of both ferro- and ferricyanide complexes was also observed by Ghosh et al. [3] in laboratory columns packed with sandy aquifer material (pH = 6.5) from an MGP site in the upper Midwest region of the United States. Figure 10.6 presents the results of a column study performed with cyanide-impacted groundwater and site aquifer material [3]. Groundwater at this site contains cyanide primarily in the form of iron–cyanide complexes (>98%), leached from Prussian Blue in soil at the MGP site. As shown in this figure, cyanide-impacted site groundwater breakthrough occurs within one pore volume, similar to a conservative tracer (chloride in this case). In addition, ferrocyanide, ferricyanide, and nickel cyanide solutions made in de-ionized water also exhibited similar transport behavior in the sandy aquifer material. Field investigations were conducted at the same MGP site in the upper Midwest region to assess the mobility of the iron–cyanide contaminated groundwater [3,61]. An extensive network of monitoring wells was installed in the thick (approximately 100 ft) unconfined aquifer underlying impacted areas within the site boundary and downgradient of the site. Groundwater samples were collected quarterly from these wells for two years, and analyzed for total cyanide, weak-acid-dissociable cyanide, and for metal–cyanide complexes by ion chromatography (selected samples). Results were used to characterize the cyanide speciation in the groundwater, and to study the transport of the cyanide. As noted above, it was found that iron–cyanide species dominated the cyanide speciation. This remained essentially constant over the two-year period of the study, reflecting the high stability of iron–cyanide species in the dark at near-neutral pH. A two-dimensional representation of the cyanide plume emanating from the MGP site source area is shown in Figure 10.7. This thin plume, which exhibited little dispersion transverse to the flow direction, was observed to move downgradient, at a rate consistent with the average linear groundwater velocity in the direction of flow. Basically, the iron–cyanide in the groundwater behaved as a nonreactive tracer. There was no apparent retardation of transport by adsorption or precipitation. This behavior was supported by the laboratory column tests described above, and by two- and three-dimensional groundwater solute transport modeling [3,61]. While the dissolved iron–cyanide was transported with the groundwater in the sand-gravel aquifer of this study site, it is not to be expected that dissolved iron–cyanide will always behave as a nonreactive tracer in groundwater. Aquifer materials with iron or aluminum oxide content will adsorb iron–cyanide species under conducive pH conditions, as noted earlier.

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FIGURE 10.7 Total cyanide plume in groundwater at a US Midwest MGP site, November 1999 (contour concentrations are mg/l). (Source: Copyright © 2001. Electric Power Research Institute, 1001301. Geochemistry, Fate and Three Dimensional Transport Modeling of Subsurface Cyanide Contamination at a Manufactured Gas Plant Reprinted with Permission.)

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10.5 SUMMARY AND CONCLUSIONS • Cyanide has been observed in soil and groundwater at many industrial sites, including electroplating facilities, aluminum and steel production plants, manufactured gas plants (MGP), and metals mining and ore heap leaching facilities. • At many cyanide-contaminated sites, including MGP and aluminum production sites, solid-phase iron–cyanide compounds are the dominant form of cyanide in soil, and dissolved iron–cyanide complexes dominate site groundwater. • Free cyanide and weak-acid-dissociable cyanides can be significant in groundwater at electroplating and metals mining and heap leaching sites, although strong metal–cyanide complexes will often still dominate because of the ubiquitous presence of iron in the soil at many sites. • Multiple physicochemical fate processes, including photodissociation, adsorption, volatilization, and precipitation govern the chemical speciation and spatial distribution of cyanide at contaminated sites. Microbiological fate processes can be important for free cyanide species, but are less relevant for metal-cyanide complexes. • Phytoextraction and phytodegradation may affect free cyanide as well as metal-complexed cyanide in the shallow subsurface at vegetated sites, depending on the types of plants present. • The mobility of cyanide in soil will be governed by precipitation–dissolution and adsorption–desorption reactions, as well as by changes in solution chemistry. Formation and dissolution of iron–cyanide solids control soluble cyanide concentrations at many sites. pH and redox potential are important parameters that control the solubility of these solids. Adsorption of cyanide species onto soils with surface reactive components (iron and aluminum oxides, clays, organic matter) can also control the mobility of cyanide in soils. • The mobility of cyanide in groundwater is dependent on soil pH, and on the types of minerals present in aquifer material. Neutral to alkaline aquifer systems with high CaCO3 content, low iron and aluminum oxide contents, low clay content, and low soil organic matter will render minimal retardation for dissolved cyanide in groundwater, while acidic soils with high clay, iron and aluminum oxide, and organic matter contents will contribute to significant retardation of dissolved cyanide in groundwater.

REFERENCES 1. USEPA, Common chemicals found at Superfund sites, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, http://www.epa.gov/superfund/resources/chemicals.htm, accessed March 22, 2005. 2. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 3. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 4. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 5. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Solubility of cyanide in contaminated soils, J. Environ. Qual., 23, 785, 1994. 6. Alesii, B.A. and Fuller, W.H., The mobility of three cyanide forms in soils, in Proceedings Residual Management by Land Disposal, Hazardous Waste Research Symposium, EPA-600/9-76-015, U.S. Environmental Protection Agency, Cincinnati, OH, 1976. 7. Dzombak, D.A., Dobbs, C.L., Culleiton, C.J., Smith, J.R., and Krause, D., Removal of cyanide from spent potlining leachate by iron cyanide precipitation, in Proceedings of WEFTEC96, Vol. 3, Part I. Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1996, p. 107.

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8. Mansfeldt, T., Gehrt, S.B., and Friedl, J., Cyanides in a soil of a former coking plant site, Z. Pflanzenernähr. Bodenk., 161, 229, 1998. 9. Meeussen, J.L., van Riemsdijk, W.H., and van der Zee, S.E.A.T.M., Transport of complexed cyanide in soil, Geoderma, 67, 73, 1995. 10. Theis, T.L., Young, T.C., Huang, M., and Knutsen, K.C., Leachate characteristics and composition of cyanide-bearing wastes from manufactured gas plants, Environ. Sci. Technol., 28, 99, 1994. 11. Weigand, H., Totschke, K.U., Mansfeldt, T., and Kogel-Knaber, I., Release and mobility of polycyclic aromatic hydrocarbons and iron cyanide complexes in contaminated soil, J. Plant Nutr. Soil Sci., 164, 643, 2001. 12. Aronstein, B.N., Maka, A., and Srivastava, V.J., Chemical and biological removal of cyanides from aqueous and soil-containing systems, Appl. Microbiol. Biotechnol., 41, 700, 1994. 13. Laha, S. and Luthy, R.G., Investigation of microbial degradation of fixed cyanide, report to the Aluminum Association by Carnegie Mellon University, Pittsburgh, PA, 1991. 14. Dhillon, J.K. and Shivaraman, N., Biodegradation of cyanide compunds by Pseudomonas species, Can. J. Microbiol., 45, 201, 1999. 15. Harris, R. and Knowles, C.J., Isolation and growth of a Pseudomonas species that utilizes cyanide as a source of nitrogen, J. Gen. Microbiol., 129, 1005, 1983. 16. Knowles, C.J., Microorganisms and cyanide, Bacteriol. Rev., 40, 652, 1976. 17. Knowles, C.J. and Bunch, A.W., Microbial cyanide metabolism, Adv. Microb. Physiol., 27, 73, 1986. 18. Knowles, C.J. and Wyatt, J.M., The degradation of cyanides and nitriles, in Microbial Control of Pollution, Fry, J.C., Gadd, G.M., Herbert, R.A., Jones, C.W., and Watson-Craik, I.A., Eds., Cambridge University Press, Cambridge, England, 1992, p. 113. 19. Staritsky, I.G., Sloot, P.H.M., and Stein, A., Spatial variability and sampling of cyanide polluted soil on former galvanic factory premises, Water, Air, Soil Pollut., 61, 1, 1992. 20. White, C.S. and Markwiese, J.T., Assessment of the potential for in situ bioremediation of cyanide and nitrate contamination at a heap leach mine in central New Mexico, J. Soil Contam., 3, 271, 1994. 21. Hayes, T.D., Linz, D.G., Nakles, D.V., and Leuschner, A.P., eds., Management of Manufactured Gas Plant Sites, Vol. 1 & 2, Amherst Scientific Publishers, Amherst, MA, 1996. 22. Lowry, H.H., Chemistry of Coal Utilization, John Wiley & Sons, New York, 1945. 23. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 24. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Dissolution behavior of iron cyanide (Prussian Blue) in contaminated soils, Environ. Sci. Technol., 26, 1832, 1992. 25. Keizer, M.G., van Riemsdijk, W.H., and Meeussen, J.L., Manganese iron cyanide as possible mineral form in contaminated non-acidic soils, Land Contam. Reclam., 3, 7, 1995. 26. Mansfeldt, T. and Dohrmann, R., Identification of crystalline cyanide-containing compound in deposited blast furnace sludge, J. Environ. Qual., 30, 1927, 2001. 27. Rennert, T. and Mansfeldt, T., Sorption of iron–cyanide complexes on goethite in the presence of sulfate and desorption with phosphate and chloride, J. Environ. Qual., 31, 745, 2002. 28. Theis, T.L. and West, M.L., Effects of cyanide complexation on the adsorption of trace metals at the surface of goethite, Environ. Technol. Lett., 7, 309, 1986. 29. Bushey, J.T. and Dzombak, D.A., Ferrocyanide adsorption on aluminum oxides, J. Colloid Interface Sci., 272, 46, 2004. 30. Rennert, T. and Mansfeldt, T., Sorption of iron cyanide complexes in soils, Soil Sci. Soc. Am. J., 66, 437, 2002. 31. Blayden, L.C., Hohman, S.C., and Robuck, S.J., Spent potliner leaching and leachate treatment, in Proc. Light Metals 1987, Denver, CO, The Minerals, Metals and Materials Society, Warrendale, PA, 1987. 32. Kimmerle, F.M., Girard, P.W., Roussel, R., and Tellier, J.G., Cyanide destruction in spent potlining, in Proc. Light Metals 1989, AIME, The Minerals, Metals and Materials Society, Warrendale, PA, 1989. 33. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. 34. EPRI, State of the science review of cyanide and its compounds at former manufactured gas plant sites, Final Report No. TR-114121, Electric Power Research Institute, Palo Alto, CA, 1999. 35. Ghosh, R.S., Nakles, D.V., Murarka, I., and Neuhauser, E.F., Cyanide speciation in soil and groundwater at manufactured gas plant (MGP) sites, Environ. Eng. Sci., 21, 752, 2004.

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36. Kjeldsen, P., Behaviour of cyanides in soil and groundwater: a review, Water, Air, Soil Pollut., 115, 279, 1999. 37. Cosgrove, J.G., Collins, R.L., and Murty, D.S., Preparation of ferrous ferricyanide (not Turnbull’s Blue), J. Amer. Chem. Soc., 95, 1083, 1973. 38. Ludi, A., Prussian blue, an inorganic evergreen, J. Chem. Educ., 58, 1013, 1981. 39. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sed. Contam., submitted, 2005. 40. Asperger, S., Kinetics of the decomposition of potassium ferrocyanide in ultra violet light, Trans. Faraday Soc., 48, 617, 1952. 41. Asperger, S., Murati, I., and Pavlovic, D., Kinetics and mechanism of the decomposition of complex cyanides of iron(II) and molybdenum(IV), J. Chem. Soc., 730, 1960. 42. Broderius, S.J. and Smith, L.L., Direct photolysis of hexacyanoferrate complexes: Proposed applications to the aquatic environment, EPA-600/3-80-003, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1980. 43. Fuller, M.W., LeBrocq, K.F., Leslie, E., and Wilson, I.R., The photolysis of aqueous solutions of potassium hexacyanoferrate (III), Aust. J. Chem., 39, 1411, 1986. 44. Gaspar, V. and Beck, M.T., Kinetics of the photoaquation of hexacyanoferrate (II) ion, Polyhedron, 2, 387, 1983. 45. Johnson, C.A., Leinz, R.W., Grimes, D.J., and Rye, R.O., Photochemical changes in cyanide speciation in drainage from a precious metal ore heap, Environ. Sci. Technol., 36, 840, 2002. 46. Kongiel-Chablo, I., Behavior of complex cyanides in natural water with high rate of contamination, Roczniki Panstwowego Zakladu Hig., 17, 95, 1966. 47. Scott Rader, W., Solujic, L., Milosavljevic, E.B., and Hendrix, J.L., Sunlight-induced photochemistry of aqueous solutions of hexacyanoferrate-(II) and -(III) ions, Environ. Sci. Technol., 27, 1875, 1993. 48. Bodek, I., Lyman, J.L., Reehl, W.F., and Rosenblatt, D.H., Environmental Inorganic Chemistry, Pergamon Press, New York, 1988. 49. Leduc, G., Pierce, R.C., and McCracken, I.R., The effects of cyanides on aquatic organisms with emphasis upon freshwater fishes, NRCC No. 19246, Associate Committee on Scientific Criteria for Environmental Quality, National Research Council of Canada, Ottawa, 1982. 50. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 51. Bushey, J.T., Modeling Cyanide Uptake by Willow for Phytoremediation. Ph.D. Thesis, Carnegie Mellon University, Pittsburgh, PA, 2003. 52. Trapp, S.A.J. and Christiansen, H., Phytoremediation of cyanide-polluted soils, in Phytoremediation: Transformation and Control of Contaminants, McCutcheon, S. and Schnoor, J., Eds., John Wiley & Sons, New York, 2003, p. 28. 53. Pettet, A.E.J. and Mills, E.V., Biological treatment of cyanides with and without sewage, J. Appl. Chem., 4, 434, 1954. 54. Dragon, T.J., Biological treatment of photo processing effluents, J. Water Poll. Control Fed., 45, 2123, 1973. 55. Oudjehani, K., Zagury, G., and Deschenes, L., Natural attenuation potential of cyanide via microbial activity in mine tailings, Appl. Microbiol. Biotech., 58, 409, 2002. 56. Mudder, T. and Whitlock, J.L., Biological treatment of cyanidation wastewaters, Miner. Metall. Proc., August, 161, 1984. 57. Boucabeille, C., Bories, A., Olliver, P., and Michel, G., Microbial degradation of metal complexed cyanides and thiocyanate from mining wastewaters, Environ. Pollut., 84, 59, 1994. 58. Thompson, L. and Gerteis, R.L., New technologies for mining waste management, biotreatment processes for cyanide, nitrates and heavy metals, in Proc. of the Western Regional Symposium on Mining and Mineral Processing Wastes, Gold Fields Mining Corporation, Golden, CO, 1990, p. 271. 59. Hommelgaard, H., Munch, S.K., Mosbaek, H., and Kjeldsen, P., Natural attenuation of iron-complexed cyanide in soil and groundwater at former gaswork sites, in Proc. Intern. FZK/TNO Conference, Contaminated Soil, Edinburgh, U.K., 1998.

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60. Roy, S.B. and Dzombak, D.A., Colloid release and transport processes in natural and model porous media, Colloids and Surfaces, A. Physicochemical and Engineering Aspects, 107, 245, 1996. 61. EPRI, Geochemistry, fate and three-dimensional transport modeling of subsurface cyanide contamination at a manufactured gas plant, Final Report No. 1001301, Electric Power Research Institute, Palo Alto, CA, 2001. 62. Ohno, T., Determination of levels of free cyanide in surface and ground waters affected by MDOT salt storage facilities, Technical Report 86C, Maine Department of Transportation, Technical Services Division, Augusta, ME, 1989. 63. Rennert, T. and Mansfeldt, T., Sorption of iron–cyanide complexes on goethite, Eur. J. Soil Sci., 52, 121, 2001.

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Cyanide in the 11 Anthropogenic Marine Environment David A. Dzombak, Sujoy B. Roy, Todd L. Anderson, Michael C. Kavanaugh, and Rula A. Deeb CONTENTS 11.1

Wastewater Discharges to Coastal Waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.1 Fate and Transport of Cyanide Species in Marine Waters . . . . . . . . . . . . . . . . . . . . . . 11.1.1.1 Onshore Tank Studies, Spiked San Francisco Bay Water . . . . . . . . . . . 11.1.1.2 Onshore Tank Studies, Mixtures of Refinery Effluent and San Francisco Bay Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.1.3 In Situ Bag Studies, Spiked San Francisco Bay Water . . . . . . . . . . . . . . 11.1.2 Ambient Water Monitoring for Cyanide in San Francisco Bay. . . . . . . . . . . . . . . . . 11.1.2.1 1989–1990 Monitoring in Suisun Bay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.2.2 1993 Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.2.3 2002–2003 Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.2.4 2003–2004 Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.1.3 Lessons from Field Observations at Wastewater Discharge Sites . . . . . . . . . . . . . . 11.2 Use of Cyanide for Capturing Live Reef Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.1 Quantity of Fish Caught Using Cyanide and the Mass of Cyanide Introduced into Coral Reefs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.2 Effects of Cyanide Use on Reefs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.3 The Future . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

210 211 211 212 212 214 214 216 216 216 218 219 219 220 221 222 222

Large quantities of anthropogenic cyanide are continuously introduced into the oceans of the world. The most significant source of cyanide input to the oceans is the atmosphere. Biomass burning introduces 1.4 to 2.9 Tg N yr−1 HCN to the atmosphere, much of which (1.1 to 2.6 Tg N yr−1 ) is estimated to be taken up by the oceans [1]. This represents a significant fraction of the estimated 30 Tg N yr−1 total input of atmospheric fixed N to the oceans [2]. There are no reported measurements of HCN/CN− in the open ocean [1]. It is estimated that free cyanide transferred to ocean water from the atmosphere is biodegraded rapidly, and has a lifetime in ocean water of a few months or less [1]. It would be difficult to detect measurable free cyanide in the oceans, as an annual accumulation of 2.9 Tg HCN as N in the top 1 m of the world oceans (approximately 3.6 × 1014 m3 ) corresponds to a concentration of 8 µg/l. This represents an unattainable upper-bound concentration, considering that free cyanide is readily biodegraded under aerobic conditions and will have a residence time of much less than one year. Thus, the actual concentration in ocean water from atmospheric input at a particular time and ocean location is likely to be considerably less than 1 µg/l, the lowest detection limit attainable with conventional analytical methods (Chapter 7). 209

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Cyanide inputs to marine waters are of greatest concern in coastal waters, where higher concentrations can be achieved, and where most recreational uses and commercial fishing occur. In coastal waters, cyanide is most commonly introduced via wastewater discharges and by the purposeful discharge involved in “cyanide fishing.” Other routes for cyanide to be introduced to coastal marine waters include spills from shoreline facilities, from ships, or loadings from rivers to which a significant upstream discharge or spill has occurred. The fate, transport, and effects of cyanide discharged to the marine environment have been studied little. It is known that free cyanide is toxic to some marine organisms at very low concentrations, significantly lower than is the case for sensitive freshwater aquatic life (Chapter 14). Accordingly, very low concentrations of cyanide are specified for marine water-quality goals. In the United States, the acute and chronic marine water-quality criteria for cyanide are 1 µg/l as free cyanide [3]. In order to be able to assess risks associated with planned or accidental discharges of cyanide that may yield such low levels in marine receiving waters, improved understanding of fate and transport of cyanide in the marine environment is needed. Such understanding is critical for determining allowable discharges to marine waters, and predicting impacts of unplanned discharges. This chapter examines what is known about the fate, transport, and effects of cyanide in marine waters by looking at the two most common scenarios for introduction of cyanide to coastal marine waters: wastewater discharges and cyanide fishing. Fate and transport of cyanide species in marine environments has not been studied systematically, but only in a general way. Focusing on the two common anthropogenic contamination scenarios provides an opportunity to identify important physical, chemical, and biological processes that influence cyanide fate in coastal marine waters. In addition, these cases are of significant national and international regulatory interest.

11.1 WASTEWATER DISCHARGES TO COASTAL WATERS Municipal and industrial wastewater discharges are important and often dominant sources of cyanide inputs to coastal waters. There are no significant natural sources of cyanide in aquatic systems (Chapter 3). Because of the high sensitivity of various marine organisms to free cyanide concentrations in water, acceptable concentrations of free cyanide in marine waters are lower than in freshwater systems. In the United States, the acute and chronic marine water-quality criteria for free cyanide are 1 µg/l compared to an acute criterion of 22 µg/l and a chronic criterion of 5 µg/l for freshwater (Chapter 14). Consequently, permissible concentrations of cyanide in discharges to coastal waters are typically very low, in some cases as low as 1 µg/l when no consideration of dilution is allowed. Concentrations at this level are at or below the limit of detection for conventional cyanide analytical methods (Chapter 7). The management of cyanide in wastewater discharges to San Francisco Bay (SFB) is a high-profile example of the challenge posed by very low regulatory limits for discharges to coastal waters [4]. Under federal and state water-quality regulations, the San Francisco Bay Regional Water Quality Control Board (SFBRWQCB) is required to establish water-quality-based effluent limits for 126 priority pollutants, of which cyanide is one. The water-quality objective for SFB is 1.0 µg/l [5], corresponding to the EPA acute and chronic marine water-quality criteria [3]. SFB receives a multitude of municipal and industrial wastewater discharges. These SFB discharges are classified for regulatory purposes as either “shallow-water” or “deep-water.” The latter are permitted an allowance for dilution in the determination of regulatory effluent limits. Shallow-water dischargers, on the other hand, are permitted no allowance for dilution. Effluent limits are set at the same levels as the ambient marine water-quality criteria, or 1 µg/l. Documenting an effluent concentration this low on a consistent basis is difficult for dischargers because the detection limits for the total and weak acid dissociable (WAD) cyanide methods commonly used in monitoring effluents are in the range of 1 to 5 µg/l [6]. Further, consistently achieving an effluent concentration this low is difficult, especially

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for municipal wastewater treatment plants in which free cyanide can be produced at low µg/l levels in chlorination and UV irradiation processes used for disinfection [7,8], or where cyanide is present in treatment plant influent and not completely destroyed in the secondary treatment process, or where cyanide-bearing scrubber water from on site sludge incineration is added to the plant influent or wastewater process stream (see Chapter 25). Due to the very low permissible amounts of cyanide in discharges to U.S. marine waters such as SFB, dischargers often seek to conduct an ecological risk assessment to provide a basis for an alternate effluent limit for cyanide (Chapter 17). An evaluation of the fate and transport of cyanide species discharged to coastal waters is a critical component in assessing the effects of cyanide exposure on local aquatic organisms. The important fate processes for cyanide in marine systems are the same as those in freshwater systems (Chapter 9): chemical transformation due to changes in solution chemistry, photodecomposition of some metal–cyanide complexes (Fe(CN)3− 6 and Fe(CN)4− 6 ), volatilization (HCN), and biodegradation (HCN). Environmental conditions such as pH, temperature, nutrient concentrations, water clarity, solar intensity, and wave activity on the surface will all affect the extent of the fate processes, as discussed in Chapter 9.

11.1.1 FATE AND TRANSPORT OF CYANIDE SPECIES IN MARINE WATERS Available information on cyanide fate and transport in marine waters is almost entirely in secondary scientific literature, in reports for studies of a particular discharge or group of discharges in a particular region. For example, S.R. Hansen and Associates of Berkeley, CA, studied the fate of metal-complexed cyanide in effluents from four petroleum refineries that discharge to San Francisco Bay [9]. This work consisted of (a) onshore, outdoor experiments in open tanks with spiked SFB water and mixtures of refinery effluent and SFB water, and (b) in-bay studies in 300-gallon translucent polyethylene bags containing refinery effluent and SFB water. Results of these studies are summarized below. 11.1.1.1 Onshore Tank Studies, Spiked San Francisco Bay Water The fate of free cyanide (KCN) and ferrocyanide (K4 Fe(CN)6 ) spiked in 30-gallon open tanks containing SFB water was studied in May 1989 by S.R. Hansen and Associates [9]. Water was obtained from Suisun Bay, which is located within the SFB estuary. No background cyanide was detected in the Suisun Bay water [9]. Two 30-gallon Nalgene tanks were filled with Suisun Bay water, individually spiked with free cyanide and ferrocyanide, and placed on a wharf on Suisun Bay. Samples were subsequently taken from each tank on a periodic basis over one week, and analyzed for total cyanide and weak acid dissociable cyanide (WAD CN). Cyanide amenable to chlorination (CATC) was also measured but CATC data are not discussed here owing to the shortcomings of this analytical method (Chapter 7). Results of the onshore tank experiments are presented in Table 11.1. As indicated in the results for the tank with the KCN spike, free cyanide was rapidly lost from the tank. Within three days the total cyanide in the tank was reduced from 20 µg/l to below the detection limit of 0.5 µg/l (modified method employed to achieve this level), likely through a combination of volatilization and biodegradation. A continuous decrease in total cyanide was also observed in the tank with the ferrocyanide spike. Total cyanide was reduced to below the detection limit in 6 days. Considering the resistance of dissolved iron-cyanide complexes to biodegradation and their susceptibility to photodecomposition (Chapter 5), these results suggest that the spiked ferrocyanide was decomposed by photolytic action, followed by biodegradation and volatilization of the resulting free cyanide. The decomposition of the ferrocyanide to free cyanide is supported by the observed significant increase in WAD cyanide

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TABLE 11.1 Results of Onshore Tank Studies with Spiked SFB Water Spike KCN KCN KCN K4 (Fe(CN)6 ) K4 (Fe(CN)6 ) K4 (Fe(CN)6 ) K4 (Fe(CN)6 )

Time (days)

CNT (µg/l)

0 1 3 0 1 3 6

20.0 11.2 <0.5 21.4 15.2 4.8 <0.5

WAD CN (µg/l) 20.2 11.2 <0.5 1.4 8.4 5.0 <0.5

Source: Data from Hansen, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990.

in the early part of the test period, followed by a steady decline in WAD cyanide due to the loss of free cyanide from the system. 11.1.1.2 Onshore Tank Studies, Mixtures of Refinery Effluent and San Francisco Bay Water The fate of WAD cyanide and strongly complexed cyanide in several refinery effluents mixed with SFB water was also investigated in onshore tank studies in July 1989 by S.R. Hansen and Associates [9]. For each refinery effluent, studies were conducted with two 30-gallon Nalgene tanks, one filled with 100% refinery effluent and one with a 50:50 volumetric mixture of refinery effluent and Suisun Bay water. Samples were subsequently taken from each tank on a periodic basis over 2 weeks, and analyzed for total cyanide and WAD cyanide. Results of the onshore tank experiments conducted with one of the refinery effluents are summarized in Table 11.2. From the time-zero data for the 100% effluent tank, it is evident that about 50% of the cyanide in the tank was present as strongly complexed cyanide, most likely iron cyanide. The remainder of the cyanide was present as free or weakly complexed cyanide. Similar to the phenomena observed in the spiked tank studies, total cyanide decreased rapidly in each of the two tanks. In the tank with the 50:50 mix of refinery effluent and SFB water, total cyanide was reduced from 67 µg/l to less than 1 µg/l within 4 days. Total cyanide in the tank with 100% effluent was reduced from 110 to 2.2 µg/l within 10 days. The data suggest that the strong cyanide complexes were decomposed, likely by photolytic action (see Chapters 5 and 9), yielding free cyanide that was removed from the system by biodegradation or volatilization. 11.1.1.3 In Situ Bag Studies, Spiked San Francisco Bay Water To investigate fate of cyanide species in coastal marine waters under environmentally realistic but controlled conditions, S.R. Hansen and Associates [9] performed studies in January 1990 with largevolume samples of spiked SFB water in translucent polyethylene bags placed in SFB, specifically in the Suisun Bay. Nylon-reinforced bags 12-ft long and 3-ft in diameter with a capacity in excess of 300 gallons were used for the experiments. The bags were attached to foam collars so that the entire 12-ft length of the bag could be maintained at the water surface after deployment. The bag and collar assemblies were lashed to wharf pilings to hold them in a relatively fixed, near-shore location. For the spiked bay water tests, two bags were filled with 300 gallons of Suisun Bay water, spiked

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TABLE 11.2 Results of Onshore Tank Studies with Refinery Effluent and SFB Water Effluent:SFB water (v:v) 100:0 100:0 100:0 100:0 100:0 100:0 100:0 100:0 50:50 50:50 50:50 50:50 50:50 50:50 50:50 50:50

Time (days)

CNT (µg/l)

0 1 2 4 5 7 9 10 0 1 2 4 5 7 9 10

110 82 44 11 4.7 4.4 2.5 2.2 67 45 21 1.6 0.5 0.8 0.6 0.6

WAD CN (µg/l) 54 37 30 12 6.2 6.2 4.5 3.7 29 25 16 2.1 1.0 1.1 1.0 0.7

Source: Data from Hansen, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990.

with KCN or K4 Fe(CN)6 to provide an initial total cyanide concentration of about 20 µg/l, sealed, and deployed. Samples were subsequently taken from each bag on a periodic basis over one week, and analyzed for total cyanide and WAD cyanide. Preliminary tests with the bags indicated the occurrence of slow leakage into and out of the bags. To account for leakage during the experiments, iodide was added to each bag as a conservative tracer at the beginning of each experiment. Samples were collected for iodide analysis at each scheduled sampling event. Measured iodide concentrations were used to correct the cyanide data for dilution effects. Results of the in situ bag experiments with spiked SFB water are presented in Table 11.3. The data for the KCN spike experiments indicate that free cyanide decreased rapidly under the ambient test conditions, with a 50% reduction in concentration within 5 days. The measured WAD cyanide concentrations were anomalously higher than the total cyanide concentrations, so an accurate halflife estimate is difficult, but the data nevertheless indicate a very rapid decrease. Since the water was contained within a closed polyethylene bag, inhibiting volatilization, it appears that the primary removal mechanism for the free cyanide was biodegradation. The results for the ferrocyanide-spiked bag presented in Table 11.3 indicate that there was about 25% reduction in total cyanide over 6 days, and some transformation of the ferrocyanide to free cyanide as WAD cyanide increased over the same period. As the tests were conducted in January, with relatively short days and overcast conditions [9], there was reduced exposure to sunlight as compared to conditions during most of the year. The rate of phototransformation of the ferrocyanide species would accordingly have been relatively low as compared to annual average conditions. Nevertheless, this short-duration experiment indicates the potential for transformation of ferrocyanide in saltwater when exposed to light under field conditions.

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TABLE 11.3 Results of In Situ Bag Studies with Spiked SFB Watera Spike KCN KCN KCN KCN KCN K4 (Fe(CN)6 ) K4 (Fe(CN)6 ) K4 (Fe(CN)6 ) K4 (Fe(CN)6 ) K4 (Fe(CN)6 )

Time (days)

CNT (µg/l)b

0 1 2 4 6 0 1 2 4 6

20.0 8.5 7.5 6.4 6.1 20.0 15.6 16.9 11.7 15.2

WAD CN (µg/l)b 19.1 15.2 13.2 11.7 8.2 2.5 2.9 13.7 7.2 8.1

a Source: Data from Hansen, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990. b Iodide tracer concentrations used to correct the data for dilution effects.

11.1.2 AMBIENT WATER MONITORING FOR CYANIDE IN SAN FRANCISCO BAY Several monitoring studies of ambient cyanide concentrations in San Francisco Bay have been conducted from 1989 to 2004, including in the vicinity of some industrial and municipal wastewater treatment plant discharges. S.R. Hansen and Associates [9] performed measurements of cyanide in Suisun Bay, near treated wastewater discharges from four petroleum refineries. Additional ambient water monitoring studies performed were: 1993 monitoring conducted by the San Francisco Estuary Institute (SFEI) under the first year of their Regional Monitoring Program (RMP) for Trace Substances [10]; 2002–2003 monitoring conducted by SFEI at three RMP monitoring stations; and 2003 and 2004 “ambient” monitoring conducted by a number of shallow-water dischargers to the SFB along their discharge gradients in addition to their effluent samples. Each of these data sets is described below. While these studies have been of limited scope and duration, the results provide insight into the sources, fate, and transport of cyanide species in coastal waters.

11.1.2.1 1989–1990 Monitoring in Suisun Bay A 10-month monitoring program was conducted in San Francisco Bay, in the vicinity of four petroleum refineries in the Suisun Bay area, from April 1989 through January 1990 [9] to assess background concentrations of total cyanide. Monthly sampling of the water column was conducted at four locations near treated wastewater outfalls from the four petroleum refineries. The treated wastewaters from the four refineries exhibited average total cyanide concentrations ranging from 6 to 123 µg/l, with WAD cyanide essentially equal to total cyanide for three of the four refineries, and a preponderance of strongly complexed cyanide in the wastewater of the fourth refinery, which also had the highest amount of total cyanide (Table 11.4). Samples were collected from both rising and ebb tide waters at each location for each sampling event. Samples were analyzed for total cyanide using a modification of the standard total cyanide analysis technique to achieve a reduced detection limit of 0.5 µg/l. The method was modified by doubling the volume of sample distilled (from 500 to 1000 ml) and by reducing the volume of the NaOH absorber solution by a factor of five (from 250 to 50 ml).

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TABLE 11.4 Average Concentrations of Total and WAD Cyanide in Treated Wastewaters of Four Petroleum Refineries, January–July 1989a Petroleum refinery

CNT (µg/l)b

A B C D

WAD CN (µg/l)b

5.6 11 8.3 123

5.9 14 13 20

a Source: Data from Hansen, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990. b Average of measurements for 26 samples from January to July 1989.

TABLE 11.5 Concentrations of Total Cyanide (µg/l) Measured at Four Locations in SFB, April 1989–January 1990a Date

Sta. A

Sta. B

Sta. C

Sta. D

4/25/89 5/16/89 6/9/89 7/13/89 8/18/89 8/26/89 9/11/89 10/31/89 12/12/89 12/23/89 1/17/90

<0.5 <0.5 <0.5 <0.5 8.0 <0.5 0.54 <0.5 <0.5 <0.5 <0.5

<0.5 <0.5 <0.5 <0.5 6.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5

<0.5 <0.5 <0.5 <0.5 6.8 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5

<0.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5 <0.5

a Source: Data from Hansen, Report by S.R. Hansen and Associates, The Joint

Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990.

Results of the 1989–1990 cyanide background concentration monitoring in SFB are presented in Table 11.5. The total cyanide measured in the 44 samples collected over the 10-month period was consistently below the detection limit of 0.5 µg/l, except for three of the four locations during the August 12, 1989 sampling event, and a single detection of 0.54 µg/l at one site during September 1989. Observed total cyanide concentrations on the August sampling day ranged from nondetectable to 8.0 µg/l. Additional samples collected two weeks later all had no detectable cyanide. The investigators were not able to explain the occurrence of the apparently anomalous data during the initial August 1989 sampling event.

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The results of the 1989–1990 SFB monitoring program conducted near four refineries suggest substantial assimilative capacity of the near-shore, relatively shallow SFB waters near the refinery treated wastewater outfalls. Average total cyanide concentrations in treated wastewater effluent as high as 123 µg/l were reduced to nondetectable (<0.5 µg/l) or near nondetectable concentrations at monitoring locations near the outfalls. 11.1.2.2 1993 Monitoring During March, May, and September 1993, the first year of the Regional Monitoring Program [10], total and WAD cyanide (Standard Methods 4500-CN-C and 4500-CN-I; see Chapter 7) were analyzed in samples obtained at 16 monitoring stations. The stations were located throughout San Francisco Bay, specifically at the deeper channels of the Bay. Samples were collected at 1-m depths during various tidal conditions. The protocols established by the RMP for Quality Assurance and Quality Control (QA/QC) were followed and analytical detection limits were 1.0 µg/l. All samples were nondetectable (<1.0 µg/l) for both total cyanide and WAD cyanide. Based on these results, it was decided to remove cyanide from the list of parameters for subsequent RMP sampling events. 11.1.2.3 2002–2003 Monitoring Results of a 2002–2003 monitoring program for a number of pollutants, including cyanide, in the San Francisco Bay are documented in a report prepared by SFEI [11]. The monitoring was conducted in response to a request from SFBRWQCB to develop information to meet state and federal requirements regarding water-quality standards and effluent limitations [12]. The purpose of the monitoring program was to allow for the determination of whether a discharger may cause, have reasonable potential to cause, or contribute to, an excursion above any applicable priority pollutant criterion or objective. Several SFB dischargers elected to meet this information collection requirement collectively, through the Bay Area Clean Water Agencies (BACWA) organization and a contract with SFEI. The 2002–2003 work consisted of three monitoring events (two in 2002, one in 2003) conducted at three historical RMP monitoring sites selected to represent the range of waters receiving discharges in the SFB area. A modified total cyanide method with a detection limit of 0.4 µg/l, or 40% of the marine water-quality criterion (WQC) for free cyanide of 1 µg/l, was used in this study. Total cyanide was only detected in one sample (at 0.5 µg/l); the remaining eight samples did not contain detectable cyanide. The study authors noted that the QA/QC results associated with the 2002–2003 cyanide results indicated there were some analytical or matrix interferences. They suggested that further work to improve analytical reliability would be needed to increase the confidence in the study findings that total cyanide concentrations throughout San Francisco Bay are consistently below the WQC of 1 µg/l. 11.1.2.4 2003–2004 Monitoring Between July 2003 and April 2004, the City of San Jose and other participants in the San Francisco Bay Shallow Water Dischargers work group voluntarily sampled “ambient” total cyanide concentrations in shallow water along their treated municipal wastewater discharge gradients in San Francisco Bay [13]. A Practical Quantitation Limit (PQL) of 0.3 µg/l was achieved utilizing a modified version of the standard colorimetric finish method for the ambient cyanide analyses (Standard Methods 4500-CN-E; see Chapter 7); a PQL of 1.0 µg/l was used for the effluent analyses. The City of San Jose also published results of its January 2003 through June 2004 cyanide monitoring associated with the San Jose/Santa Clara Water Pollution Control Plant (WPCP) [14]. Collective results of the Shallow Water Dischargers work group and City of San Jose sampling are presented in Table 11.6,

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TABLE 11.6 Average and Range of Concentrations of Total Cyanide in SFB Water near Five Shallow Water Discharges, 2003–2004 Shallow water treated municipal wastewater discharger Fairfield-Suisuna Las Gallinasa Petalumaa San Jose/Santa Clarab Sonoma County Water Agencya

Sampling period

No. of ambient water stations

No. of water samples

CNT (µg/l) range

CNT (µg/l) mean

February 2004 April 2004 February–April 2004 July 2003–June 2004 February–April 2004

8 4 2 13 7

8 8 6 172 10

0.4–1.6 <0.3–2.7 0.5–0.9 <0.3–59 0.3–2.9

1.0 1.6 0.7 1.3 1.0

a Source: Data from LWA, Memo to S. Moore and P. Bossak of the San Francisco Bay Regional Water Quality Control Board

from T. Grovhoug and A.K. Brinton of LWA, June 16 (revised November 29), Larry Walker Associates, Lafayette, CA, 2004. b Sources: Data from both LWA, Memo to S. Moore and P. Bossak of the San Francisco Bay Regional Water Quality Control

Board from T. Grovhoug and A.K. Brinton of LWA, June 16 (revised November 29), Larry Walker Associates, Lafayette, CA, 2004; and CSJ, Cyanide attenuation study report: investigation into the fate and transport of cyanide in the wastewater treatment process at the San Jose/Santa Clara Water Pollution Control Plant and the determination of a receiving water attentuation factor for cyanide in Lower South San Francisco Bay, City of San Jose, CA, 2004.

where it may be seen that ambient levels of total cyanide ranged from a low of less than 0.3 µg/l (nondetectable) near the Las Gallinas and San Jose/Santa Clara discharges to a high of 59 µg/l near the San Jose/Santa Clara discharge. (The 59 µg/l value was associated with a single, atypical monitoring event; the next highest ambient concentration recorded at one of the San Jose/Santa Clara ambient sampling locations was 6 µg/l.) Mean total cyanide concentrations for the five shallowwater dischargers ranged from 0.7 µg/l near the Petaluma discharge to 1.6 µg/l near the Las Gallinas discharge. In comparison, effluent concentrations of total cyanide during the July 2003 to April 2004 time period ranged from a low of less than 1.0 µg/l (nondetectable; 5 of the 7 dischargers) to 63 µg/l at the San Jose/Santa Clara WPCP. (This 63 µg/l value is associated with the same atypical monitoring event as discussed above. The next highest San Jose/Santa Clara WPCP effluent cyanide value was 8 µg/l; the next highest value among all dischargers was 10.6 µg/l, at Palo Alto.) Mean total cyanide effluent concentrations for the six dischargers ranged from 0.4 µg/l (Las Gallinas effluent) to a high of 6.8 µg/l (Palo Alto). From four of the dischargers with both effluent and ambient monitoring results, a limited number of conclusions can be made regarding effluent and ambient cyanide. First, the Fairfield–Suisun and Petaluma data sets are insufficient with respect to indicating a potential relationship between effluent and ambient cyanide concentrations, because all data are very close to their respective detection/reporting limits and were deemed insufficient to develop mean effluent values, as indicated in Table 11.7. Second, two other data sets suggest different relationships between effluent and ambient concentrations: the Las Gallinas data indicate slightly higher ambient concentrations in terms of both range and mean as compared to effluent data, while Sonoma County Water Agency data indicate the opposite. Reaching conclusions regarding the relationship of effluent to ambient concentrations from these data is difficult because (1) the time period for the effluent data was July 2003–April 2004, a larger time window than most of the ambient data sets, and (2) both data sets consist primarily of data very close to their respective detection/reporting limits. The City of San Jose, which has conducted more intensive effluent and ambient cyanide monitoring than the other dischargers noted in this section, has concluded that significant attenuation of cyanide occurs between their effluent discharge and ambient monitoring stations near their discharge.

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TABLE 11.7 Average and Range of Concentrations of Total Cyanide in Effluent from Seven SFB Treated Municipal Wastewater Discharges, 2003–2004 Treated municipal wastewater discharger Fairfield-Suisuna Las Gallinasa Palo Altoa Palo Altoa,b Petalumaa San Jose/Santa Clarac Sonoma County Water Agencya Sunnyvalea

Sampling period

Number of effluent samples

CNT (µg/l) range

CNT (µg/l) mean

July 2003–April 2004 July 2003–April 2004 July 2003–April 2004 July 2003–April 2004 July 2003–April 2004 January 2003–June 2004 July 2003–April 2004 July 2003–April 2004

6 18 8 8 3 70 11 16

<1.0–1.1 <1.0–1.3 4.2–10.6 1.9–5.8 <1.0–1.2 1.5–63 <1.0–5.8 <1.0–2.8

IDd 0.4 6.8 4.8 IDd 3.5 1.9 1.1

a Source: Data from LWA, Memo to S. Moore and P. Bossak of the San Francisco Bay Regional Water Quality Control

Board from T. Grovhoug and A.K. Brinton of LWA, June 16 (revised Novemeber 29), Larry Walker Associates, Lafayette, CA, 2004. b Indicates unpreserved samples; all other results in table reflect preserved samples. c Source: Data from CSJ, Cyanide attenuation study report: investigation into the fate and transport of cyanide in the wastewater treatment process at the San Jose/Santa Clara Water Pollution Control Plant and the determination of a receiving water attentuation factor for cyanide in Lower South San Francisco Bay, City of San Jose, CA, 2004; low end of range also defined by several results recorded as <5 µg/l. d Insufficient data to quantify result.

Based on its 2003 and 2004 cyanide monitoring data [14], the City of San Jose developed “attenuation factors” for cyanide ranging from 1.3 to 7.4, averaging 2.9. These factors represent ratios between plant effluent and ambient cyanide concentration values and were developed without incorporating the singly atypically high monitoring event (which would have resulted in an attenuation factor of 19) [14].

11.1.3 LESSONS FROM FIELD OBSERVATIONS AT WASTEWATER DISCHARGE SITES Based on observations made in San Francisco Bay in the vicinity of treated wastewater discharges from four petroleum refineries with relatively high average total cyanide concentrations (from 5.6 to 123 µg/l), it appears that cyanide in the discharges undergoes attenuation to very low levels (less than or close to 0.5 µg/l) in the vicinity of discharges, even when complexed cyanide is present in the discharges. Similar studies in the vicinity of treated municipal wastewater discharges to SFB also generally showed decreases but were more difficult to interpret as the total cyanide concentrations in both the municipal wastewater effluents (<1 to 10 µg/l) and in the bay waters (<0.3 to 2.9 µg/l) were close to analytical detection limits. While dilution of the discharges with SFB water has a role in reducing concentrations, there is field evidence that other fate processes have significant effects. Some innovative in situ bag studies conducted in near-shore areas of SFB demonstrated that ferrocyanide can decompose rapidly by photolysis under surface water solar exposure, and that free cyanide biodegrades or is volatilized rapidly. The field observations coupled with the 3− in situ bag studies suggest that photolysis (for dissolved iron cyanide, Fe(CN)4− 6 and Fe(CN)6 ), biodegradation (HCN), and volatilization (HCN) indeed have a significant influence on the fate of

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cyanide discharged to coastal waters, as discussed in Chapter 9. However, additional studies are needed in other locations representing a range of wastewater types and coastal environments to form general conclusions about assimilative capacity of coastal waters for cyanide.

11.2 USE OF CYANIDE FOR CAPTURING LIVE REEF FISH Another prevalent source of cyanide release to coastal marine waters occurs during the practice of cyanide fishing in coral reefs in Southeast Asia. Cyanide fishing is a technique that has been used for several decades for capturing aquarium fish and, more recently, for capturing live food fish. This method of fishing typically involves fishermen diving near coral reefs and squirting fish with a relatively high concentration of sodium cyanide solution or using cyanide-poisoned bait. Cyanide interferes with oxygen uptake in fish, as discussed in greater detail in Chapter 14, and following exposure, the fish are stunned and easy to capture alive. Fish that survive the initial cyanide exposure are able to metabolize and excrete the cyanide while they are held in intermediate holding tanks, and become a valuable export commodity. Aside from the effects of cyanide on fish that enables their capture, controlled studies have also demonstrated the adverse effects of this practice on coral reefs, such as discoloration and death of coral polyps [15–17]. Cyanide fishing practiced repeatedly in the vicinity of reefs leads to their destruction. Cyanide fishing also carries human risks, such as exposure of fishermen to the chemical, and to fish consumers who eat fish that may contain cyanide residues. Human health risks of cyanide are discussed in Chapters 13 and 16. Cyanide fishing appears to have originated in the Philippines in the late 1950s to early 1960s, originally as a means to capture ornamental aquarium fish, and has gradually spread to other nations in Southeast Asia such as Indonesia, Thailand, Taiwan, Maldives, and others [18]. The practice is illegal in most countries, although enforcement is relatively lax in all but a few. Reliable data on the quantity of fish caught using this technique and the areas where it is commonly used are difficult to obtain. However, studies by various nongovernmental groups indicate that the practice is still common in the Philippines, is spreading rapidly in Indonesia, and is possibly being practiced in Papua New Guinea and other islands in the South Pacific [19]. The reef fish species that are most commonly targeted by cyanide fishermen for the live food fish trade are groupers, wrasses, and coral trout, though not all cyanide fishing is targeted to individual species [20]. There are no analogous data on the species targeted for the aquarium fish trade, in part because a wider range of species and a larger number of fish are caught. However, the tonnage of the aquarium trade and its economic value are estimated to be significantly lower than the food fish trade. The major destinations of the food fish and aquarium fish are different. Food fish is primarily consumed in the Far East with Hong Kong and Southern China widely reported to be key export markets [21,22]. Aquarium fishes are exported much more widely, with North America being a major destination [18].

11.2.1 QUANTITY OF FISH CAUGHT USING CYANIDE AND THE MASS OF CYANIDE INTRODUCED INTO CORAL REEFS Official data on the live fish trade, caught with and without cyanide, are generally unreliable with only a few countries reporting any statistics. Researchers studying the problem of cyanide fishing have used various indirect methods to estimate the contribution of cyanide fishing, and also the quantity of cyanide released into coral reef environments. One approach to estimating the quantity of cyanide-caught fish is to assume that it constitutes a fraction of the total live fish trade. Based on recent data from Hong Kong, it is estimated that about 30,000–35,000 tons of live food fish are imported [21,23,24]. The Hong Kong trade is thought to be 65 to 80% of the global food fish

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trade [22], which is estimated to be about 38,000–54,000 tons. Using a conservative estimate of about 20% of live fish being caught with cyanide, based on a random analysis of traded live fish in the Philippines in 1996 [18], it can be estimated that approximately 12,000–14,000 tons of live food fish are caught using cyanide annually. This is a conservatively low estimate because the use of cyanide in other regions involved in this trade may be more common than in the Philippines. Cyanide is also used in the capture of live aquarium fish, and Rubec et al. [25] reported detection of 20% of aquarium fish with cyanide residues. Based on data from the Philippines, the number of live aquarium fish traded may be much greater than food fish, because the individual fish are much smaller. Nevertheless, the total mass of aquarium fish exported is perhaps a tenth of the food fish exports [18]. Estimates of cyanide release into reefs, much like the estimates of the live fish trade, are, by the very nature of the data, approximate and perhaps speculative at this stage. Because cyanide is a common industrial chemical with uses in mining, electroplating, and other industries, it is not possible to use data on cyanide manufacture or import quantities to estimate what is being used for capturing fish. Estimates of the quantities of cyanide released into reefs vary widely. Using the quantity of cyanide individual fishermen use as a starting point, studies have estimated that between 150,000 and 500,000 kg of sodium cyanide may be used annually in the Philippines [26], and 320,000 to 640,000 kg in Indonesia [27] for collection of food and aquarium fish from reefs. These numbers may be compared with the total cyanide imports of 7 million kg into a country such as the Philippines (1995 data, reported in Barber and Pratt [18]). Although the Philippines and Indonesia are thought to be the largest users of cyanide for fishing, the adoption of this approach in other regions of the world may mean that over a million kg of sodium cyanide is directly released into reefs each year.

11.2.2 EFFECTS OF CYANIDE USE ON REEFS Although a large number of anecdotal reports on the loss of corals due to cyanide exposure are available (e.g., documented by Johannes and Riepen [19] and Cesar [27]), most of these reports are not in scientific literature and little can be inferred from them about the concentrations and time periods of exposure that led to harmful effects. A limited number of controlled laboratory studies have shown the range of concentrations and the mechanisms through which cyanide affects corals [16,17]. Jones and Steven [17] isolated pieces of two common species of corals from the Great Barrier Reef and subjected them to cyanide exposure at different concentrations and for different durations. The corals were exposed to concentrations of 2×10−1 , 2×10−2 , 2×10−3 , and 2×10−4 M NaCN for durations of 1, 5, 10, 20, or 30 min. Following the exposure, the corals were examined visually for effects such as mortality or discoloration (the common term for this phenomenon is “coral bleaching”) and the symbiotic photosynthetic algae growing on the corals, the zooxanthellae, were characterized by their density (algae per unit area) and pigment concentration. For most coral species, the observed colors are in fact the colors of the zooxanthellae. Besides providing color, zooxanthellae photosynthesis also provides a food source for coral polyps, and their health is vital to the health of the coral community. At the highest concentration, 2 × 10−1 M NaCN, exposure for durations greater than 1 min led to coral mortality, and in every instance corals were discolored. At 2 × 10−2 M, no mortality was observed, but discoloration and loss of zooxanthellae was common, especially over 5 min of exposure. At 2 × 10−3 M and at exposures of 10 min or more, one of the coral species showed greater sensitivity, some discoloration, and some zooxanthellae loss. There was no visible effect at the lowest studied concentration of 2 × 10−4 M; however, at 20 min or longer exposure, zooxanthellae density showed a significant decline. Thus, for at least one of the coral species studied, there was an adverse effect at every cyanide concentration that was used. In a later paper, Jones and Hoegh-Guldberg [16], using fragments of a coral common in the Indo-Pacific region in a controlled laboratory study, demonstrated that the mechanism of coral discoloration was caused by cyanide interfering with the photosynthetic processes of the zooxanthellae. An adverse effect on zooxanthellae density was seen at concentrations as low as 10−5 M cyanide following

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exposure for 3 h in the presence of light. The same levels of cyanide exposure in darkness did not cause adverse effects on zooxanthellae. Cervino et al. [15] conducted laboratory studies of cyanide exposure on eight species of corals from the Indo-Pacific region. Coral fragments were exposed for 1 to 2 min in a tank containing seawater with cyanide at either 1 × 10−3 , 2 × 10−3 , 0.6 × 10−2 , or 1.2 × 10−2 M, removed, washed, and placed in a tank with cyanide-free seawater. The coral fragments were then observed for responses from 1 to 12 weeks. Immediately upon exposure at all concentrations, the corals secreted mucus and retracted their tentacles. Over time, some individuals of all coral species exhibited mortality, even at the lowest dose level used in the study. Discoloration and reduced zooxanthellae density was also observed for most coral species. These changes were significantly different from the controls that had not been exposed to cyanide. Together this body of literature provides support for the fact even when brief cyanide exposure is not immediately lethal, major adverse effects occur over longer time frames with the potential for severe impacts on coral communities. These effects are consistent with field observations that when coral reefs are damaged by cyanide, death of the organisms is not immediate, but occurs over a period of weeks or months [19]. No field data exist on the concentrations of cyanide that result from fishermen’s use of cyanide for fishing. The concentrations that may result from a typical use of cyanide can be approximated as follows: assume the concentration in the squirt bottle is 2 M and that 0.33 l of solution is squirted into a volume that rapidly disperses over a volume of 1 m3 , resulting in an initial concentration of approximately 0.7 × 10−3 M. Although this concentration will diminish rapidly due to dilution in this volume, the brief contact of reef organisms with high concentrations of cyanide, based on the controlled studies described above, appears to be sufficient to cause permanent long-term impacts. Moreover, cyanide fishing is not always practiced with squirt bottles for capture of specific fish. There have been reports of fishermen sometimes dumping entire drums of cyanide into shallow reef communities [19,20], which would produce very high concentrations in the vicinity of the discharge. According to Johannes and Riepen [19], the objective in this practice is not to obtain live fish, but simply to maximize the catch of dead fish with a minimum of effort. The practice appears to be used when the weather is rough, conventional fishing is difficult, and the fishermen need the money [19].

11.2.3 THE FUTURE The demand for live reef fish for food is expected to grow. Available information collected by nongovernmental agencies studying reef fishing indicates that cyanide fishing continues to be prevalent in Southeast Asia, especially in Indonesia [22], a country that is home to a large fraction of the world’s coral reefs. Aside from developed countries such as Australia, the Philippines stand out as a developing nation where regulatory mechanisms have had some effect in controlling the use of this method, although even here cyanide fishing has not been completely eradicated [18,25,26]. It is also possible that because cyanide has been used for a much longer time in the Philippines, fewer locations with adequate stocks of commercially valuable fish species are available compared to the less exploited reefs in Indonesia. Most countries that either export or import live fish have made cyanide fishing illegal. However, weak enforcement typically prevails. In the future, international trade may be restricted due to considerations of several endangered species of fish that are targeted by cyanide fishermen, such as some groupers and Napoleon wrasse. These fish are slow-moving and territorial, and their populations may be depleted relatively quickly. Also, testing of fish after capture may be used more extensively in the future to detect fish caught with cyanide, and thereby restrict their trade. A laboratory program in the Philippines has already had some success in identifying cyanide-caught fish. Although cyanide fishing occurs in remote areas, many of which are economically deprived, a relatively sophisticated infrastructure must exist for the captured fish to be transported alive to their final markets, which are in more economically developed countries. Over time, there may be greater control of this practice in the final markets than in the source areas [22].

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Although reef damage by cyanide has been documented to some extent, it is important to consider that there are other factors that may have equally, if not more, significant adverse effects on reefs. These include potential increases in ocean temperatures due to global warming, pollution from other toxins and nutrient runoff, especially near populated and agricultural areas, and blast fishing, where explosives are used to blast corals and capture the stunned fish [24,27,28]. A long-term assessment of the damage caused by cyanide fishing must also consider these other hazards.

11.3 SUMMARY AND CONCLUSIONS • HCN in the atmosphere from global biomass burning is transferred into the oceans of the world at the rate of 1.1 to 2.6 Tg N yr−1 . This represents a significant fraction of the estimated 30 Tg N yr−1 total input of atmospheric fixed nitrogen to the oceans. • There are no reported measurements of HCN/CN− in the open oceans. Free cyanide transferred to ocean water from the atmosphere likely undergoes rapid biodegradation. • Cyanide inputs to marine waters are of greatest concern in coastal waters, where higher concentrations can be achieved, and where most recreational uses and commercial fishing occur. • Two important sources of cyanide discharges in coastal marine waters are wastewater discharges and cyanide fishing around coral reefs. • In the United States, permissible amounts of cyanide in discharges to coastal waters are typically very low, in some cases as low as 1 µg/l when no consideration of dilution is allowed. • The limited available data for cyanide in coastal waters in the vicinity of wastewater discharges indicate that free cyanide species are rapidly biodegraded and volatilized, and that strongly complexed cyanide, which usually is in the form of iron cyanide species, undergoes fairly rapid photolytic decomposition to free cyanide. Additional systematic studies are needed to confirm this for a broader range of discharge types and environmental conditions, however. • Cyanide fishing, which is practiced primarily in Southeast Asia, involves fishermen diving near coral reefs and squirting fish with a relatively high concentration of sodium cyanide solution. Nontargeted fishing involving the dumping of whole drums of sodium cyanide into shallow reef areas also occurs. • Cyanide fishing practiced repeatedly in the vicinity of coral reefs leads to damage or even destruction of the reefs. • Cyanide fishing also leads to increased human risks, through exposure of the fishermen to the chemical, and to fish consumers who eat fish containing cyanide residues. • Experimental studies of cyanide effects on corals show that even at concentrations as low as 10−5 M, adverse effects on symbiotic algae on corals (zooxanthellae) can be detected following brief periods of exposure.

REFERENCES 1. Li, Q., Jacob, D.J., Bey, I., Yantosca, R.M., Zhao, Y., Kondo, Y., and Notholt, J., Atmospheric hydrogen cyanide (HCN): biomass burning source, ocean sink? Geophys. Res. Lett., 27, 357, 2000. 2. Duce, R.A., Liss, P.S., Merrill, J.T., Atlas, E.L., and Buat-Menard, P., The atmospheric input of trace species to the world oceans, Global Biogeochem. Cycles, 5, 193, 1991. 3. USEPA, Ambient water quality criteria for cyanide — 1984, EPA-440/5-84-028, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 1984. 4. Deeb, R.A., Dzombak, D.A., Theis, T.L., Ellgas, W., and Kavanaugh, M.C., The cyanide challenge, Water Environ. Technol., 15, 35, 2003.

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5. SFBRWQCB, Water quality control plan (basin plan) for the San Francisco Bay basin, San Francisco Bay Regional Water Quality Control Board, California Environmental Protection Agency, http://www.waterboards.ca.gov/sanfranciscobay/basinplan.htm, accessed: March 8, 2005. 6. Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Evaluation and testing of analytical methods for cyanide species in municipal and industrial contaminated waters, Environ. Sci. Technol., 37, 107, 2003. 7. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 8. Zheng, A., Dzombak, D.A., Luthy, R.G., Kavanaugh, M.C., and Deeb, R.A., The occurrence of cyanide formation in six full-scale publicly owned treatment works, Water Environ. Res., 76, 101, 2004. 9. Hansen, Studies in support of alternate cyanide effluent limits for four San Francisco Bay area refineries, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990. 10. SFEI, San Francisco estuary regional monitoring program for trace substance, 1993 Annual Report, San Francisco Estuary Institute, San Francisco, CA, 1994. 11. SFEI, San Francisco Bay ambient water monitoring interim report, Report for San Francisco Bay Regional Water Quality Control Board, San Francisco Estuary Institute and Bay Area Clean Water Agencies (BACWA), San Francisco, CA, 2003. 12. SFBRWQCB, Requirement for monitoring of pollutants in effluent and receiving water to implement new statewide regulations and policy, Memorandum to permitted wastewater dischargers, San Francisco, CA, San Francisco Bay Regional Water Quality Control Board, 2001. 13. LWA, Basin plan assistance Task 5: summary of shallow water discharger effluent and ambient concentration data for cyanide (revised draft), Memo to S. Moore and P. Bossak of the San Francisco Bay Regional Water Quality Control Board from T. Grovhoug and A.K. Brinton of LWA, June 16 (revised November 29), Larry Walker Associates, Lafayette, CA, 2004. 14. CSJ, Cyanide attenuation study report: investigation into the fate and transport of cyanide in the wastewater treatment process at the San Jose/Santa Clara Water Pollution Control Plant and the determination of a receiving water attentuation factor for cyanide in Lower South San Francisco Bay, City of San Jose, CA, 2004. 15. Cervino, J.M., Hayes, R.L., Honovich, M., Goreau, T.J., Jones, S., and Rubec, P.J., Changes in zooxanthellae density, morphology and miotic index in hermatupic corals and anemones exposed to cyanide, Mar. Poll. Bull., 46, 573, 2003. 16. Jones, R.J. and Hoegh-Guldberg, O., Effect of cyanide on coral photosynthesis: implications for identifying the cause of coral bleaching and for assessing the environmental effects of cyanide fishing, Mar. Ecol.: Prog. Ser., 177, 83, 1999. 17. Jones, R.J. and Steven, A.L., Effects of cyanide on corals in relation to cyanide fishing on reefs, Mar. Freshwater Res., 48, 517, 1997. 18. Barber, C.V. and Pratt, V.R., Sullied seas: strategies for combating cyanide fishing in southeast Asia and beyond, World Resources Institute, Washington, D.C., 1997. 19. Johannes, R.E. and Riepen, M., Environmental, economic and social implications of the live reef fish trade in Asia and the Western Pacific, The Nature Conservancy, Jakarta, Indonesia, 1995. 20. Stevens, W.K., A food fad’s ripple effect on reefs of Pacific: cyanide, The New York Times, October 31, 1995, p. A1. 21. Chan, P., The industry perspective: wholesale and retail marketing aspects of the Hong Kong live reef food fish trade, SPC Life Reef Fisheries Information Bulletin, May, 2000, p. 3. 22. Graham, T.R., A collaborative strategy to address the live reef food fish trade, Report #0101, Pacific Coastal Marine Program, The Nature Conservancy, Honolulu, HI, 2001. 23. Lau, P.P.F. and Parry-Jones, R., The Hong Kong trade in live reef fish for food, TRAFFIC East Asia and World Wildlife Fund for Nature, Hong Kong, 1999. 24. Mous, P.J., Pet-Soede, L., and Erdmann, M., Cyanide fishing on Indonesia coral reefs for the live food market: what is the problem? SPC Coastal Fisheries Programme, http://www.spc.int/ coastfish/News/LRF/7/LRF7-07.htm, September, 2003.

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25. Rubec, P.J., Cruz, F., Pratt, V.R., Oellers, R., and Lallo, F., Cyanide-free net-caught fish for the marine aquarium trade, SPC Live Reef Fish Fisheries Information Bulletin, May, 2000, p. 28. 26. Rubec, P.J., Cruz, F., Pratt, V.R., Oellers, R., McCullough, B., and Lallo, F., Cyanide-free net-caught fish for the marine aquarium trade, Aquarium Sci. Conserv., 3, 37, 2001. 27. Cesar, H., Economic analysis of Indonesian coral reefs, Work in Progress Series Report, World Bank, Environment Department, Washington, DC, 1996. 28. McManus, J.W., Reyes, R.B., and Nanola, C.L., Effects of some destructive fishing methods on coral cover and potential rates of recovery, Environ. Manage., 21, 69, 1997.

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12 Cyanide Cycle in Nature Rajat S. Ghosh, Stephen D. Ebbs, Joseph T. Bushey, Edward F. Neuhauser, and George M. Wong-Chong CONTENTS 12.1

Global Output of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.1.1 Natural Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.1.2 Anthropogenic Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2 Physical, Chemical, and Biological Processes that Influence Environmental Fate of Cyanide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.1 Cyanide Compound Dissociation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.2 Cyanide Compound Solubility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.3 Metal-Cyanide Dissociation by UV Light Irradiation . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.4 Hydrogen Cyanide Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.5 Cyanide Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2.6 Biological Degradation of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3 Why does Cyanide not build up in Nature? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.1 Biotic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.3.2 Abiotic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4 Cycling of Anthropogenic Cyanide in Nature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5 Combined Natural and Anthropogenic Cyanide Cycle. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.6 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

226 226 226 227 227 228 228 228 228 229 229 230 230 231 233 235 235

Chapters 1, 3, and 4 describe the production of cyanide via natural and anthropogenic sources. A range of plants and organisms produce cyanide (Chapter 3). Natural or anthropogenic cyanide deposited or discharged to soil or surface waters can be transported by water (Chapters 9 and 10), and be subject to biodegradation, adsorption, or volatilization in these systems. In the absence of natural sinks, cyanide would accumulate in the air, soil, and water (both surface and ground) environments, proving catastrophic for plant and animal life. However, accumulation does not occur in spite of significant localized discharge episodes (e.g., the January 2000 Baia Mare spill in Romania — see Chapter 1). Natural cyanide sinks exist, including microorganisms capable of degrading and assimilating cyanide (Chapter 6), certain classes of plants capable of up-taking cyanide (Chapter 24), and soils with high anion exchange capacities and organic content that can immobilize the movement of cyanide (Chapter 5). Although of less import than the carbon or nitrogen cycle, there is a natural cyanide cycle and this cycle is linked with the nitrogen cycle. Cyanide in nature is continually produced, transported, and transformed in the environment and anthropogenic cyanide released to the environment enters this natural cycle. This chapter summarizes the effects of various biological, physical, and chemical fate processes on natural and anthropogenic cyanides and examines the components of the cyanide cycle 225

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and their effects on naturally produced cyanide, anthropogenic cyanide, and combined natural and anthropogenic cyanide.

12.1 GLOBAL OUTPUT OF CYANIDE The global output of cyanide comes from natural and anthropogenic processes, which are summarized here. Detailed discussions of these sources of cyanide are provided in Chapters 3 and 4.

12.1.1 NATURAL SOURCES Natural sources of cyanide and the mechanisms involved in the production of these cyanogenic compounds are discussed in Chapter 3. These natural sources, comprising plants, microbes, and animals, can be summarized as follows: • Plants: All plants produce some level of cyanide; 2650 species representing some 130 families of plants produce significant levels of cyanide. • Microbes (algae, bacteria, and fungi): • Algae: Three species have been identified with capabilities to produce cyanide: Chorella vulgaris, Anacystis nidulans, and Nostoc muscorum. The latter two species belong to the blue-green class of algae. • Bacteria: Cyanogenesis in bacteria is limited to Chromobacterium violaceum and certain pseudomonad species (i.e., P. aeroginosa, P. aureofaciens, P. chlorophis, and P. fluorescens). • Fungi: Cyanogenesis has been identified in numerous species of fungi of the following genera: Actinomycetes, Basidomycetes, Clitocybe, Marasmius, Pholiota, Polyporus, and Tricholoma. Many of these organisms are pathogenic to plants. • Animals: Cyanogenesis in animals is limited to certain classes of arthopods: • Chilopods (centipedes): seven species of the known 3,000 species. • Diplopods (millipedes): 46 species of the known 7,500 species. • Insecta (insects): 68 species of the 750,000 known species. By far, plant sources of cyanogenic compounds are the most significant natural sources in terms of quantity of cyanide produced, as shown in Figure 12.1, and global economic and environmental impacts. Sorghum is a widely used forage in livestock agriculture, and cassava (Manihot esculente) is a food staple for over 500 million people in Asia, Africa, South America, and the Caribbean. Events of human health disasters from the ingestion of improperly processed cassava are well documented and discussed in Chapter 3.

12.1.2 ANTHROPOGENIC SOURCES Cyanide is an important feedstock to industry and is a commercial commodity. As discussed in Chapter 4, in 2003 some 2.6 million tons of cyanide were produced and used globally. Catalytic synthesis of hydrogen cyanide from ammonia and methane accounts for a major portion while by-product recovery from acrylonitrile manufacturing accounts for the remaining portion. Some of the uses of cyanide and cyanide compounds include: synthetic fibers and plastics; gold mining; agricultural herbicides, fumigants, and insecticides; dyes and pigments; animal feed supplements; chelating agents for use in water treatment, and specialty chemicals and pharmaceuticals [1]. Cyanide is also produced in “incidental” amounts in the course of many manufacturing operations: aluminum smelting; coke, iron, and steel making; petroleum refining; municipal waste sludge incineration; and coal gasification. In the past, industrial manufacturing operations discharged cyanides to the environment in significant quantities, but current discharges are highly regulated and thus much more controlled.

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227

0.07–0.3 mg/g

Free cyanide concentration

Cyanide Cycle in Nature

FIGURE 12.1 Distribution of naturally occurring cyanide in the environment. (Source: ATSDR, U.S. Department Health and Human Services, Agency for Toxic Substances and Disease Registry, Toxicological profile for cyanide (update), 1997.)

An example of a former industrial operation disposing significant quantities of cyanide to the environment is a manufactured gas plant (MGP). These coal gasification facilities were built in cites and towns in much of the United States and Europe in the mid-to-late 1800s to provide gas for heating and lighting. Hydrogen cyanide and hydrogen sulfide were removed from the product gas by reaction with iron filings in scrubber boxes (iron oxide boxes). The cyanide formed solid phase ferric-ferrocyanide (FFC) or Prussian Blue, in these iron oxide boxes. On-site disposal of spent iron oxide boxes from MGP operations [2] has resulted in release of dissolved iron-cyanide, thiocyanate, and trace amounts of nickel and free cyanide [3–7]. Historic on-site disposal of cyanide-bearing spent potliner materials from aluminum smelting operations also has resulted in the release of dissolved cyanide in groundwater at these sites. Cyanide in spent potliner leachate is primarily in the form of iron-cyanide complexes [8–10] because of the reactivity of cyanide and the abundance of iron in soil. Sodium ferrocyanide is used as anticaking agent in road salt, giving the salt a blue color. The use of cyanide-treated road salt has resulted in the presence of iron-cyanide in roadway runoff [11].

12.2 PHYSICAL, CHEMICAL, AND BIOLOGICAL PROCESSES THAT INFLUENCE ENVIRONMENTAL FATE OF CYANIDE Details of physicochemical and biological processes that affect the fate of cyanide species in the environment are provided in Chapters 2, 5, and 6. A summary of these processes is provided here to outline the basic properties that affect the cyanide cycle in the environment.

12.2.1 CYANIDE COMPOUND DISSOCIATION Dissolved cyanide species, including free cyanide (HCN, CN− ) and various metal-cyanide complexes, can dissociate to various extents, the degree of which depends upon the pH, the temperature, and redox potential of the solution. Cyanide complexed with alkali metal (e.g., Na, K) and alkaline earth metal (e.g., Ca, Mg, and Ba) is readily dissociable in aqueous solutions, yielding cyanide ion

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(CN− ) and the corresponding cation. Certain transition metal-cyanide complexes such as those with Cu, Ag, Zn, Cd, Ni, and Hg, are considered weak acid dissociable (WAD) complexes, since they dissociate under weak acid pH (pH ∼ 4–5) conditions. Finally, there are complexes with certain transition metals (e.g., Au, Fe, Pt, Pd, and Co) that are highly resistant to dissociation and that only dissociate under strong acid conditions (pH ∼1–2) and elevated temperatures (100◦ C). In fact, for complete dissociation of cobalt-cyanide complexes ultraviolet light irradiation is included with the low pH and high temperature conditions.

12.2.2 CYANIDE COMPOUND SOLUBILITY The solubility of various solid phase cyanide compounds is discussed in detail in Chapters 5 and 10. Cyanide compounds in disposed wastes from industrial activities, like FFC, have varying degrees of solubility when contacted with water. FFC is relatively insoluble under near neutral and acidic pH conditions (pH ≤ 6). However, as pH of the water is increased from near neutral to alkaline conditions and oxygen is introduced, increased solubility occurs with the release of dissolved ferrocyanide ([Fe(CN)6 ]4− ). Other simple cyanide compounds, like alkali or alkaline earth metal-cyanide salts which occur in mine tailings, are readily soluble in water.

12.2.3 METAL-CYANIDE DISSOCIATION BY UV LIGHT IRRADIATION Chapter 5 presents details on the effects of ultraviolet light irradiation on dissolved metal-cyanide complexes. The most common and important effect of ultraviolet light irradiation on any cyanide species in the environment is the photodissociation of dissolved metal-cyanide complexes, especially the iron compounds. Numerous laboratory and a few field studies have shown that ferrocyanide complex dissociates under sunlight [4,12–15]. The rate of photochemical dissociation is greatly dependent on various environmental factors, which include but are not limited to pH, free cyanide content of the solution, sunlight intensity, temperature, turbidity, and depth of the water column. The free cyanide resulting from photodissociation could possibly be undetectable or short lived [16,17]. Detailed information is provided in Chapters 5 and 9.

12.2.4 HYDROGEN CYANIDE VOLATILIZATION At pH below 9.2 and 20◦ C any aqueous solution containing free cyanide is dominated by dissolved hydrogen cyanide which is a volatile species. The rate of volatilization of HCN from water to air, that is, from aqueous HCN to HCN gas, is governed by Henry’s law for equilibrium partitioning, by physical conditions including water turbulence and temperature, and by the chemical quality of the water. Hydrogen cyanide has moderate volatility, and given enough air:water surface area and time a significant portion of mass of HCN in water can be transferred to the gas phase.

12.2.5 CYANIDE PRECIPITATION Dissolved metal-cyanide complexes and free cyanide ion can precipitate under acidic conditions in the presence of excess iron. Other metals, like Cu and Zn, can also form precipitates with cyanide. However, these precipitates often do not form in the environment because of the abundance of iron in water and soils, resulting in the preferential formation of iron-cyanide solids. Formation of metalcyanide precipitates in the presence of excess iron has a significant role in controlling mobility of dissolved cyanide in groundwater and hence the cyanide cycle in nature. Detailed information on metal-cyanide precipitation is provided in Chapter 5.

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12.2.6 BIOLOGICAL DEGRADATION OF CYANIDE Free and WAD cyanides are biologically degradable. The mechanisms of these transformations and assimilation are discussed in detail in Chapter 6. Expressions for both aerobic and anaerobic transformation reactions are presented in Table 6.2. With strongly complexed metal (e.g., iron) cyanides biological transformation will occur to the extent that these compounds dissociate and yield free cyanide.

12.3 WHY DOES CYANIDE NOT BUILD UP IN NATURE? Cyanide recycling occurs in nature. There are various processes by which cyanide is retained and sometimes transformed in environmental compartments such as water, soil, air, and plants, and at the same time there are source processes by which cyanide is generated and released into the same compartments. Some common sinks for natural and anthropogenic forms of cyanide are: • Various species of plants that use CN− as a source of nitrogen • Plant uptake of metal-cyanide complexes and subsequent conversion to free cyanide or incorporation into the plant N cycle • Bacterial degradation of free cyanide in soil • Fungal degradation of cyanide • Adsorption in soils There are organisms that can produce and release cyanide into the environment (Chapter 3). For example, herbivores that feed on cyanogenic plants can detoxify the cyanide, or die, decay, and return cyanide (CN− ) or nitrogen to the soil where biodegradation and transformation can occur. Figure 12.2 depicts a generalized cyanide cycle that considers the recycling of cyanide from natural sources, principally cyanogenic glycosides from plants, with no input from anthropogenic sources. As shown in Figure 12.2, the entire cycle can be considered to comprise several key compartments (A, B, C, D, and E) and associated processes, with cycling of cyanide from Plant cyanogenic glycosides A

Herbivores (cyanogenic glycosides hydrolyzed to HCN)

Cyanide as part of ethylene synthesis

B CN–/HCN in dead animals converted to Feces of insects and damaged N compounds leaves release SCN and free cyanide C

Cyanogenic N incorporated into plant amino acids Ground surface

D

A B

E D

Insects feeding on leaves

C

Soil bacteria and fungi convert SCN and free cyanide (CN–) to ammonia to nitrate

Plant source of nitrogen (residual free cyanide, nitrate and ammonia)

E

Groundwater

FIGURE 12.2 Natural cyanide cycle.

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A → B → C → D → E → A being a typical pathway. In this cycle, A represents an input of plant cyanogenic glycosides that is subsequently consumed by insects and other animals (B). Upon consumption, cyanide is converted to SCN− that is returned to soil via fecal material or animal decay, while dead animals not consumed by predators can release any free cyanide to soil (C). In the soil, fungi and bacteria convert free cyanide to ammonia, followed by conversion of ammonia to nitrate (D). Any residual free cyanide and nitrogen compounds formed from free cyanide can be released to groundwater (E) or utilized by plants (A), thus completing the cycle. Biotic and abiotic sinks and sources of cyanide involved in the generalized cycle are discussed in this section.

12.3.1 BIOTIC PROCESSES The presence of cyanide in nature has been postulated to be a component of a cyanide minicycle [18,19] involving cyanide producing (cyanogenic) organisms and organisms that assimilate cyanide as a source of carbon and nitrogen for growth. An example of cyanide generation involves the cyanogenic plant birdsfoot trefoil (Lotus corniculatus), larvae of the five spotted burned moth (Zygaena trifolii), snails (Helix aspersa), and an ichneumonid wasp (Apantales zygaebarun) [20]. Production of cyanide by birdsfoot trefoil deters/limits herbivory by this snail, with the snail feeding only on tissues with low cyanide concentrations. In contrast, the moth larvae preferentially consume highly cyanogenic material and sequester the cyanide for their own defense. This sequestered cyanide deters predation of the larvae, except by A. zygaebarum, which can internally detoxify the cyanide present in the moth larvae. Another example of cyanide generation involves sorghum forage. Sorghum (e.g., Sorghum bicolor) is a cyanogenic plant that produces cyanogenic glycosides as a defensive compound. Sorghum can become infected with the pathogen Gloeocercospora sorghi, a fungus capable of detoxifying cyanide. The fungus, having the ability to metabolize cyanide, feeds upon the cyanide released by sorghum, assimilating this molecule as a source of carbon and nitrogen to continue growth and advance the infection [19]. Infected sorghum, or sorghum plants that have been cut and are decomposing, will continue to release cyanide into the soil. The cyanide released represents a potentially valuable source of carbon and nitrogen that opportunistic organisms will utilize. An illustration of cyanide cycling in nature occurred in the spring of 2001 when 500 foal and fetus deaths occurred on central Kentucky farms [21]. Agronomists reported heavy infestation of eastern tent caterpillars in Kentucky during that period. Eastern tent caterpillars, ferocious leaf eaters, ate cyanogenic leaves of black cherry trees during the spring season. One hypothesis maintains that the caterpillars left cyanide-laced feces on the surrounding grass. The mares suffered cyanide poisoning from incidentally ingesting the feces of the tent worms, resulting in miscarriages. Another hypothesis is that there was prodigious germination of cherry seedlings in the pastures and that these seedlings were consumed directly by the horses. The cause of this event has not been fully explained but the evidence suggests a central role of the cyanide minicycle. A major sink for cyanide in soil is biodegradation, where the microbes can covert free and WAD cyanide to ammonia and subsequently to nitrate [22–25]. Free cyanide not biodegraded could either be taken up by plants [26], where enzymatic processes incorporate CN− into amino acids [27,28], sorbed onto soil [29,30], or transported via groundwater. Another major sink in the soil would be organic matter and metal complexation immobilization of free and soluble cyanide (Chapters 5 and 10).

12.3.2 ABIOTIC PROCESSES Under normal pH conditions (4 < pH < 9) in natural waters, free cyanide is mostly present as aqueous HCN. Dissolved HCN as well as certain metal-cyanide complexes, like ferro- and ferricyanide complexes, can partition to and be retained on particles through adsorption. HCN adsorbs onto organic

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particles and mineral particles with organic content. The cyanide anion and metal-cyanide anions adsorb onto oxide minerals such as iron oxides, aluminum oxides, and clays with high anion exchange capacities [7,29,31]. The extent of adsorption is highly dependent on solution conditions, especially pH. Conversely, quartz minerals do not have significant affinity for free cyanide or dissolved ironcyanide species [3]. Adsorption of cyanide species onto minerals and minerals with organic matter is reversible, with desorption inducible by changing solution conditions, especially pH.

12.4 CYCLING OF ANTHROPOGENIC CYANIDE IN NATURE Cyanide removal processes active within the natural cyanide cycle also act on anthropogenic cyanide. One example is solid phase FFC, which was often land disposed on site during past industrial practices at former MGP sites. Leachates from this disposed material may contain iron-cyanide complexes and small amounts of other weak acid dissociable complexes depending on geochemical conditions [3]. These soluble cyanide species can be taken up by plants [26,32,33] or can be transported along with the groundwater as shown in Figure 12.3. Selenocyanate disposed in refinery sludge can also be taken up and metabolized by plants [34]. Cycling of anthropogenic cyanide, depicted in Figure 12.3, occurs in the same manner as the natural cyanide cycle discussed above, except here the cyanide input is from an anthropogenic source. This anthropogenic cyanide input could be transported via groundwater, could be taken up by plants or could be metabolized by soil microorganisms. Similar recycling of anthropogenic cyanide generated by the mining industry (e.g., mine tailings) is discussed by Smith and Mudder (1991). As shown in Figure 12.4, reproduced from Smith and Mudder (1991), water bearing cyanide in a tailings pond is in contact with tailings solids and sediment, and the atmosphere. Hence, the fate and transport of aqueous cyanide is represented by a cyanide cycle encompassing a complex set of chemical reactions involving free cyanide radical and various metal-cyanide complexes. Biodegradation of free and WAD cyanide dominates in the tailings sediment, while metal-cyanide complexation, precipitation, biological oxidation, photolysis, Plant cyanogenic glycosides A

Herbivores (cyanogenic glycosides hydrolyzed to HCN)

Cyanide as part of ethylene synthesis

Insects feeding on leaves

B CN–/HCN in dead animals Feces of insects and damaged converted to N compounds leaves release SCN and free cyanide C

Cyanogenic N incorporated into plant amino acids Ground surface

FFC used as fill

A

Soil bacteria and fungi convert SCN and D free cyanide (CN–) to ammonia to nitrate

leachate B

E

Plant source of nitrogen (residual free cyanide, [Fe(CN)6]3–/4–, nitrate and ammonia)

D

C

Groundwater

FIGURE 12.3 Recycling of anthropogenic cyanide via natural cyanide cycle.

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E

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Fe4 [Fe(CN)6]3 Prussian Blue

Precipitation

[Fe(CN)6]4–

CN–

Fe

2+

Fe3+

C

moisture/ r ain/fog

[Fe(CN)6]3–

Ultraviolet light

H C N/ C N – (extremely dilute)

CN– Partial biological oxidation

Soil/ surface water

S0

CN – Mx–y

[M (CN)x ]y –

SCN–

Anaerobic biological activity

B

NH3, CO2, CH4, H2S

Biological oxidation

NH3, HCO3– , HSO4–

Tailings p ond

HCO3– , NH4+ Biological oxidation

HCOO– , NH4+

Biological oxidation

Hydrolysis

The atmosphere

Diffusion/dispersion

Stratospheric disappearance

where, M = Ni, Cu, Zn

H C N/ C N –

A

HCN

B

HCN Volatilization

A

C

FIGURE 12.4 Cyanide cycle in a mine tailing pond/atmosphere system. (Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991, with permission.)

• Hydrolysis • Biological oxidation • Plant nutrient • Animal metabolism

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233

UV

HCN(g) Volatilization

CNin

Uptake

CN–

Biodegradation CNout

Fe(CN)64– Soil sorption Me–CN Solid

Adsorption on Fe(OH)3(s)

Soil CNout <<< CNin

FIGURE 12.5 Cyanide fate mechanisms in a constructed wetland. (Source: Bushey, J.T., Ph.D. thesis, Carnegie Mellon University, 2003, with permission.)

and volatilization are the primary fate mechanisms operating in the tailings pond water. Figure 12.4 shows that cyanide recycling occurs between the tailings pond water and the atmosphere (A → B → C → A). For example, a portion of free cyanide in the water column (A) can volatilize as gaseous HCN (B) into the atmosphere where it can be absorbed into moisture droplets (C) and reenter the tailings pond via precipitation. The other portion can be biologically degraded to ammonia, carbon dioxide, and nitrate. The natural processes that are responsible for removal and recycling of cyanide in the environment can be exploited in engineered systems to treat cyanide from industrial discharges. An example of one such engineered system is a constructed wetland to manage cyanide-impacted surface water or groundwater (Figure 12.5). Several studies, two related to the evaluation of wetlands for treatment of cyanide discharges, and others related to cyanide phytoremediation, are presented in Chapter 24. As shown in Figure 12.5, similar to the tailings pond, both biological and physicochemical processes control the fate of cyanide in a wetland environment. These include: • Transport processes (advection and dispersion) • Photolysis (dissociation of iron/other metal-cyanide complexes to free cyanide via UV irradiation) • Volatilization of free cyanide into the atmosphere • Microbiological degradation of free cyanide and weakly-complexed cyanide • Plant uptake and assimilation of free and metal-complexed cyanide • Rhizosphere mediated degradation of cyanide species • Adsorption of free and metal-cyanide complexes onto sediments • Precipitation/dissolution of iron and other metal-cyanide complexes.

12.5 COMBINED NATURAL AND ANTHROPOGENIC CYANIDE CYCLE An example of combined cycling of natural and anthropogenic cyanide is presented in detail in Figure 12.6. This figure shows that anthropogenic cyanide of various forms, as solid ferric-ferrocyanide (FFC) and yellow Prussiate of Soda (Na4 Fe(CN)6 ), and dissolved forms

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A

Plant synthesis of cyanogenic glycosides

Rain/precipitation Herbivores (cyanogenic glycosides hydrolyzed to HCN)

Cyanide as a N source in plants Dissolved cyanide uptake via roots (as free and complex cyanide)

CN–/HCN via plant decay

C

Vadose Zone B

E D

C

Cyanide released into atmosphere as HCN Photolysis & volatilization

CN–/HCN in dead animals converted to N compounds

Industrial CN, e.g., FFC, YPS A

B

CN in discharges

Soil bacteria and fungi convert CN– to CO2, Surface water NO3– or NH3 D body

leached into GW as CN– , Dissolved cyanide transport in groundwater SCN–, [M(CN)x ]y– Plant source of N, NO3– ,NH3 E Aquifer

CN–/[M(CN)x ]y– sorbed/precipitated in soil

FIGURE 12.6 Example of a combined cyanide cycle.

TABLE 12.1 Literature Rates for Various Environmental Fate Processes in the Cyanide Cycle Fate processes

Species

Process rate

Source

Phyto-uptake

Fe(CN)6 4− CN−

2 mg as CN/kg(FW)-day 8 mg as CN/kg(FW)-day

[26]

Cyanogen glycoside production

CN−

600 mg CN/kg-day (max rate)

[35]

Biodegradation

Fe(CN)6 4−

[36]

CN−

No ubiquitous organisms; process rates using specific engineered organisms: 530,000 mg CN/day-kg cells 600,000 mg CN/day-kg cells

[22]

Photodissociation

Fe(CN)6 4−

0.09 to 9 mg as CN/L-h

[37]

Adsorption on aluminum oxide

Fe(CN)6 4−

5 mg as CN/kg solid-day

[38]

(e.g., SCN− , CN− and various metal ([M(CN)x ]y− ) cyanides) can be incorporated into the natural cyanide cycle by the operation of the various fate mechanisms described previously in this chapter. Natural cyanide is recycled via the A → B → C → D → E → A pathway, where plant cyanogenic glycosides (A) are consumed by insects and other animals (B), which eventually return to soil via fecal material or animal decay (C). In the soil, microbial degradation converts a portion of the free cyanide to nitrate and ammonia (D); the remaining free cyanide combines with that from anthropogenic sources (E) that can be released to groundwater, can be adsorbed and/or precipitated in soil and/or removed by plant uptake (A), thus completing the cycle. Laboratory generated removal rates of free and complexed iron-cyanide via different fate processes are summarized in Table 12.1. These process rates depend on many factors. For example, plant

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uptake, microbial degradation, and adsorption depend on the cyanide species, microbial population, aquatic chemistry, and site specific physicochemical conditions which can vary considerably from site to site. Nonetheless, the rates presented in Table 12.1 provide relative order-of-magnitude removal rates for these fate processes and they help to assess the importance of individual fate processes operating in the combined cyanide cycle (Figure 12.6).

12.6 SUMMARY AND CONCLUSIONS • Free cyanide occurs naturally in the environment, being produced by a variety of plants and organisms, and is recycled through a combination of physical, chemical, and biological processes. • Both plants and microbes can utilize cyanide as a source of nitrogen for growth or incorporate it into their metabolic system. • Free cyanide and metal-cyanide complexes are used as raw materials for the manufacture of many useful products and are also produced as by-products from many manufacturing operations. These anthropogenic activities have introduced additional cyanide into the natural cycle. • Anthropogenic cyanide is readily incorporated into the natural cyanide cycle, and natural mechanisms for cyanide removal from the environment can be exploited for management of anthropogenic cyanide.

REFERENCES 1. Pesce, L.D., Cyanides, in Kirk–Othmer Encyclopedia of Chemical Technology, Vol. 7, John Wiley & Sons, New York, 1993. 2. Hayes, T.D., Linz, D.G., Nakles, D.V., and Leuschner, A.P., Eds., Management of Manufactured Gas Plant Sites, Vols. I and II, Amherst Scientific Publishers, Amherst, MA, 1996. 3. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 4. Meeussen, J.L., Keizer, M.G., and de Haan, F.A.M., Chemical stability and decomposition rate of iron-cyanide complexes in soil solutions, Environ. Sci. Technol., 26, 511, 1992. 5. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Solubility of cyanide in contaminated soils, J. Environ. Qual., 23, 785, 1994. 6. Meeussen, J.L., van Riemsdijk, W.H. and van der Zee, S.E.A.T.M., Transport of complexed cyanide in soil, Geoderma, 67, 73, 1995. 7. Theis, T.L. and West, M.L., Effects of cyanide complexation on the adsorption of trace metals at the surface of goethite, Environ. Technol. Lett., 7, 309, 1986. 8. Blayden, L.C., Hohman, S.C., and Robuck, S.J., Spent potliner leaching and leachate treatment, in Proceedings of Light Metals 1987, Denver, CO, The Minerals, Metals and Materials Society, Warrendale, PA, 1987, p. 663. 9. Dzombak, D.A., Dobbs, C.L., Culleiton, C.J., Smith, J.R., and Krause, D., Removal of cyanide from spent potlining leachate by iron-cyanide precipitation, in Proceedings of WEFTEC96, Vol. 3, Part I. Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1996, p. 107. 10. Robuck, S.J. and Luthy, R.G., Destruction of iron-complexed cyanide by alkaline hydrolysis, Water Sci. Technol., 21, 547, 1989. 11. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron-cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 12. Broderius, S.J. and Smith, L.L., Direct photolysis of hexacyanoferrate complexes: proposed applications to the aquatic environment, EPA-600/3-80-003, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1980. 13. Burdick, G.E. and Lipschuetz, M., Toxicity of ferro- and ferricyanide solutions to fish and determination of the cause of mortality, Trans. Am. Fish. Soc., 78, 192, 1948.

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14. Fuller, M.W., LeBrocq, K.F., Leslie, E., and Wilson, I.R., The photolysis of aqueous solutions of potassium hexacyanoferrate (III), Aust. J. Chem., 39, 1411, 1986. 15. Gaspar, V. and Beck, M.T., Kinetics of the photoaquation of hexacyanoferrate (II) ion, Polyhedron, 2, 387, 1983. 16. Laszlo, Z. and Dombi, A., Oxidation of [Fe(CN)6 ]4− and reduction of [Fe(CN)6 ]3− in VUV-irradiated aqueous solutions, Chemosphere, 46, 491, 2002. 17. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 18. Allen, J. and Strobel, G.A., The assimilation of H14 CN by a variety of fungi, Can. J. Microbiol., 12, 414, 1966. 19. Thatcher, R.C. and Weaver, T.L., Carbon–nitrogen cycling through microbial formamide metabolism, Science, 192, 1234, 1976. 20. Jones, D.A., Selective eating of the acyanogenic form of the plant Lotus corniculatus by various animals, Nature, 193, 1109, 1962. 21. USAT, Cherry tree leaves killed foals, scientists conclude, USA Today, July 13, 2001. 22. Dhillon, J.K. and Shivaraman, N., Biodegradation of cyanide compounds by Pseudomonas species, Can. J. Microbiol., 45, 201, 1999. 23. Harris, R. and Knowles, C.J., Isolation and growth of a Pseudomonas species that utilizes cyanide as a source of nitrogen, J. Gen. Microbiol., 129, 1005, 1983. 24. Knowles, C.J., Microorganisms and cyanide, Bacteriol. Rev., 40, 652, 1976. 25. Knowles, C.J. and Bunch, A.W., Microbial cyanide metabolism, Adv. Microb. Physiol., 27, 73, 1986. 26. Ebbs, S.D., Bushey, J.T., Poston, S., Kosma, D., Samiotakis, M., and Dzombak, D.A., Transport and metabolism of free cyanide and iron-cyanide complexes by willow, Plant Cell Environ., 26, 1467, 2003. 27. Akopyan, T.N., Braunstein, A.E., and Goryachenkova, E.V., β-Cyanoalanine synthase: purification and characterization, Proc. Natl. Acad. Sci. USA, 72, 1617, 1975. 28. Warrilow, A.G.S. and Hawkesford, M.J., Separation, subcellular location and influence of sulphur nutrition on isoforms of cysteine synthase in spinach, J. Exp. Bot., 49, 1625, 1998. 29. Alesii, B.A. and Fuller, W.H., The mobility of three cyanide forms in soils, in Proceedings Residual Management by Land Disposal, Hazardous Waste Research Symposium, EPA-600/9-76-015, U.S. Environmental Protection Agency, Cincinnati, OH, 1976. 30. Rennert, T. and Mansfeldt, T., Sorption of iron–cyanide complexes on goethite, Eur. J. Soil Sci., 52, 121, 2001. 31. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sed. Contam., submitted, 2005. 32. Samiotakis, M. and Ebbs, S.D., Possible evidence for transport of an iron-cyanide complex by plants, Environ. Pollut., 127, 169, 2004. 33. Trapp, S., Larsen, M., and Christiansen, H., Experimente zum verbleib von cyanid nach aufnahme in pflanzen, Umweldt Schad Forsch, 13, 29, 2001. 34. de Souza, M.P., Pickering, I.J., Walla, M., and Terry, N., Selenium assimilation and volatization from selenocyante-treated Indian mustard and muskgrass, Plant Physiol., 128, 625, 2002. 35. Halkier, B.A., Scheller, H.V., and Moller, B.L., Cyanogenic glucosides: the biosynthetic pathway and the enzyme system involved, in Cyanide Compounds in Biology, Evered, D. and Garnety, S.F., Eds., John Wiley & Sons, Chichester, UK, 1988, p. 49. 36. Dursun, A.Y., Calik, A., and Aksu, Z., Degradation of ferrous (II) cyanide complex ions by Pseudomonas fluorescens, Process Biochem., 34, 901, 1999. 37. Johnson, C.A., Leinz, R.W., Grimes, D.J., and Rye, R.O., Photochemical changes in cyanide speciation in drainage from a precious metal ore heap, Environ. Sci. Technol., 36, 840, 2002. 38. Bushey, J.T., Modeling cyanide uptake by willow for phytoremediation, Ph.D. thesis, Carnegie Mellon University, Pittsburgh, PA, 2003.

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13 Human Toxicology of Cyanide Joseph L. Borowitz, Gary E. Isom, and David V. Nakles CONTENTS 13.1 13.2

Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sites of Action. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2.1 Brain . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2.2 Heart . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2.3 Liver and Kidney . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3 Symptoms and Signs Caused by Cyanide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.1 Acute. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.2 Aftereffects of Acute Exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.3 Chronic Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.4 Generation and Metabolism of Cyanide in Body Tissues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5 Mechanisms of Cyanide Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.1 Acidosis. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.2 Neurotransmitter Release . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.3 Alteration of Calcium Balance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.4 Oxidative Stress . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.5 Cyanide and Mitochondria. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.6 Nerve Cell Death Pathways and Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5.7 Neuronal Proteins that Influence the Toxic Effects of Cyanide . . . . . . . . . . . . . . . . . 13.6 Cyanide Antidotes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.7 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Hydrocyanic acid (HCN) was first isolated by the Swedish chemist Karl Scheele in 1782. He also became the first victim of cyanide gas when he accidentally broke a vial in the laboratory and died from vapor inhalation [1]. Many instances of murder, suicide, and accident involving cyanide have occurred through the years [2]. Cyanide is now considered a potential weapon for mass destruction in the United States for use in terrorist attacks [3,4]. This chapter describes how people can be exposed to free cyanide (HCNg , HCNaq , CN− aq ), symptoms they might experience due to actions on the brain and heart which are the main targets for cyanide, and biochemical and molecular mechanisms of action of cyanide in the human body. The chapter also reviews how small amounts of cyanide can be generated in the body. In addition, antidotal treatments for people exposed to cyanide are covered. It is important to note that this chapter focuses on actions of free cyanide, and all discussion of cyanide herein refers to cyanide in the forms listed above or the cyanide ion (CN− ) in salts that dissociates readily in aqueous solution. Metal cyanide salts and dissolved species generally exert toxicity only insofar as they release free cyanide. Cyanide complexed strongly to metals (e.g., in potassium ferricyanide) is much less toxic than simple cyanide salts, since little free cyanide is released from strong complexes even in aqueous solution [5]. 237

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13.1 EXPOSURE Cyanide is a widely used industrial chemical (see Chapter 4), and accidental human exposure to HCNg , simple cyanide salts (e.g., NaCN, KCN, CuCN), and water solutions of these salts represents a public health concern [6]. Exposure may occur in a variety of occupational, environmental, and dietary conditions. The most current National Occupational Exposure survey estimates that more than 150,000 workers are regularly exposed to cyanide compounds [6] by inhalation, ingestion, and dermal exposure. These workers are primarily in the metal processing, electroplating, plastic, and chemical synthesis industries. Environmental exposure is associated with ground water contaminated by cyanide, such as effluents from landfills and waste discharge from industrial processes using cyanide (see Chapter 16). Since combustible materials can release cyanide, the most frequent source of cyanide poisoning is smoke inhalation that is associated with a high mortality and postintoxication morbidity [7]. In the 1997 Toxicological Profile for Cyanide [6], environmental contamination, routes of exposure, animal experiments, and cases of human poisoning were included. Small amounts of cyanide can be obtained from drinking water (∼2 µg/day), foods (e.g., lima beans 0.1–3 µg/g, cereal grains 0.001–2.45 µg/g), and ambient air (∼3.8 µg/day). United States commercial cigarettes each contain 10–400 µg and can be a major source of cyanide. The Toxicological Profile [6] mentioned a case of industrial exposure to cyanide involving workers extracting silver from x-ray film [8]. This case illustrates the effects of high cyanide exposures in humans. Cyanide chelates silver from film and silver can be recovered by this process. However high concentration free cyanide solutions are needed and workers were exposed to about 15 ppm of HCN in the air each workday for an average of 11 months. Ingestion of foods and beverages in work areas as well as skin contact with cyanide solutions compounded the problem. Thirty-six workers were surveyed after they had left the silver extraction facility (average time elapsed from exposure 10.5 months). Symptoms experienced by the workers during their tenure at the facility were surveyed as well as symptoms which persisted after termination of employment. One employee actually died of cyanide poisoning, indicating the intensity of exposure of some workers. The four most common complaints were headache, dizziness, nausea/vomiting, and a bitter or almond taste. Also reported were sleep disturbances, ringing in the ears, skin rash, cough, and numbness of the extremities. Five of the workers reported loss of consciousness again reflecting the severity of the intoxication. Incidence of symptoms decreased after termination of employment but headache and eye irritation persisted in nearly half the individuals. The important publication by Blanc et al. [8] documents effects in humans after high doses of cyanide.

13.2 SITES OF ACTION 13.2.1 BRAIN Neural tissue uses large quantities of oxygen and cannot function when oxygen levels are low. As cyanide blocks oxidative metabolism, it is not surprising that neurons are a prime target for cyanide. Neural effects are prominent with both the acute and chronic effects of cyanide. Not all brain areas are equally susceptible to the action of cyanide. Even the older literature is in good agreement regarding brain areas most easily damaged by cyanide. Experiments involving rats, cats, or monkeys given cyanide repeatedly for 1 to 25 days consistently show lesions in “somatosensory cortex, hippocampus, and basal ganglia” [9–12]. Experimental acute cyanide poisoning in dogs showed lesions in cerebral cortex, basal ganglia, thalamus, and cerebellum if animals survived 3 h or more [13]. Rats given KCN subcutaneously (1.43 mg/kg once weekly for 22 weeks) showed lesions in cortex and cerebellum [14]. Neurons outside the brain, however, are not sensitive to the effects of cyanide. Thus brain areas controlling skeletal muscle activity (basal ganglia, cerebellum), memory (hippocampus), and cognitive ability (cortex) are more sensitive to cyanide’s actions.

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Acute sublethal exposure to cyanide has behavioral effects. Mild intoxication (20 to 40 ppm HCN for several hours) causes fear, nausea/vomiting, vertigo, headache, tachypnea, tachycardia, disturbed vision, excitement, and inability to concentrate [15,16]. Obviously many neural systems are activated by exposure of humans to moderate amounts of cyanide. Animal studies employing sublethal doses of cyanide also show behavioral deficits. Ability of mice to cling to a wire mesh screen is diminished by cyanide [17]. Cynomologous monkeys exposed to 60 ppm HCN for 30 min showed no deficit in skeletal muscle twitch but a decrease in brain electrical activity was noted. Also a state of semiconsciousness and severe central nervous system depression was apparent even though the animals were restrained [18]. Sublethal doses of cyanide decrease swimming performance in guinea pigs [19]. Doses of 4 mg/kg NaCN given subcutaneously ( 23 the lethal dose in 50% of the animals [LD50 ]) increased time needed to complete the swim test when the trial was begun 2 to 64 min after injection. Swim time was normal when the test was begun at 128 min after cyanide, reflecting time needed for recovery from the dose of cyanide. At 2 mg/kg NaCN had no influence on total swim time, but significantly increased the time needed to reach the first of two light beams. This occurred between 2 and 8 min after injection at a time when the pigs presumably showed little gross incoordination. Thus behavioral deficits can be produced by cyanide in doses of 13 to 23 of the LD50 . Doses of 2 to 3 (but not 1) mg/kg NaCN also decreased conditioned shock avoidance of guinea pigs and rats tested immediately after injection [19]. Animals did not jump to a safe area of the cage after a “conditioning” signal. There was however a corresponding increase in the number of escapes from shock suggesting that motor performance was unimpaired. It appears that these doses of cyanide disrupt cognitive mechanisms so that the animals are not able to process information in order to respond to the conditioned stimulus but when the shock is initiated, they have no trouble moving to the safe area. An important study by Pavlakovic et al. [20] illustrates the effectiveness of low doses of cyanide in the brain and the variability of cyanide action in different brain parts. They showed 4 mg/kg intraperitoneally in mice increased levels of a protein (Fos) in cortex and cerebellum but decreased Fos in hippocampus. Fos turns on genetic mechanisms, is a marker of neuronal activation, is detectable in all brain areas, and can be altered by many physiological and pharmacological stimuli. These authors also showed that a very low dose of cyanide (0.5 mg/kg) significantly decreased Fos in hippocampus but not in other brain areas. Lack of a uniform response to cyanide reflects the complexity of cyanide’s neurotoxic action (Figure 13.1). Recent experiments show that hearing loss may be caused by cyanide especially with simultaneous noise exposure. Fechter et al. [21] treated rats with 30 ppm of HCN for 3.5 h. Noise exposure alone impaired hearing and caused outer hair cell loss when tested 4 weeks later. HCN alone caused no hearing impairment or outer hair cell loss, but when combined with noise a potentiation of hearing impairment and outer hair cell loss was noted. In another study rats given 7 mg/kg of KCN intraperitoneally showed impairment of cochlear function with a preferential effect on high tones [22]. Specialized sensory cells can be damaged by the action of cyanide especially when these sensory neurons are activated. The studies cited above generally employed high doses of cyanide. Unfortunately effects of low levels of cyanide on human and animal behavior have not been examined. Most likely low doses have subtle effects and only careful and perhaps prolonged experiments are likely to yield positive results. Yet there is a need for such studies to be done.

13.2.2 HEART The heart is an important site for the toxic action of cyanide. The enzyme that is a primary target for cyanide (cytochrome oxidase) is inhibited to a greater extent in the heart than in brain, kidney, or liver [1]. Thus cyanide can produce heart failure [23] and also irregular heart beats (arrhythmias) in animals [24] and man [25]. A profound slowing of the heart rate (bradycardia) followed by ventricular

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250

**

**

0.5 mg/kg KCN 4 mg/kg KCN

Fos (% control)

200

150

100

*

**

50

0 Cortex

Hippocampus

Cerebellum

Brain stem

FIGURE 13.1 Effect of low (0.5 mg/kg, ip) and high (4 mg/kg, ip) doses of KCN on Fos expression in rats 60 min after cyanide administration. Data are expressed as percentage of control Fos levels (saline-injected animal) for each experimental group. Each bar represents mean ±SE of data obtained from 5 animals (*P < 0.05, **P < 0.01). Changes in Fos protein in cells is a measure of cell activation or adaption. This IEG (immediate early gene) product can change in response to toxic insults. Note that cyanide has different effects in different brain areas. Cortex and cerebellum were affected only by the high dose but a significant change occurred in hippocampus at a dose 5% of the LD50 . By contrast, no change at any dose was detected in brain stem. (Source: Pavlakovic, G., et al., Neurochem. Res., 19, 1289, 1994, with permission.)

fibrillation (grossly uncoordinated contraction) was seen in the cat after a high dose (2 mg/kg NaCN intravenously) [25]. Cyanide alters electrical activity of the heart. The “action potential” which occurs with each heart beat is shortened by cyanide and this effect is counteracted by glucose [26]. The mechanism is simple and involves decreased availability of adenosine triphosphate (ATP) following blockade of energy metabolism by cyanide. Lack of ATP impairs potassium ion channels in cell membranes and potassium efflux increases. Some believe that activation of this potassium current is an important adaptive mechanism to protect the heart from decreased oxygen availability when blood flow or oxygen supply is limited [27]. Other electrocardiographic changes caused by cyanide have been attributed to reflex effects related to low blood pressure, acidosis, direct effects of cyanide on the heart, and effect of anoxia related to inhibition of breathing [1]. Cyanide also releases neurotransmitters and adrenomedullary hormones partly by blocking calcium efflux and thus increasing intracellular calcium [28]. The heart is flooded with neurotransmitters but because of impaired metabolic processes cannot respond properly and loss of function with arrhythmias occurs. Baskin et al. [29] isolated guinea pig hearts to study effects of cyanide in both electrically paced and unpaced preparations and thus eliminated any reflex or humoral or breathing problems. In the first minute of cyanide (0.5 mM) perfusion, an increased force of contraction was noted in paced hearts and this effect was blocked by propranolol indicating release of norepinephrine was involved. Cyanide reduced both rate and contractile displacement in unpaced hearts. The decrease in heart rate occurred prior to the effect on contraction. Generation of an electrical pulse appears to be more sensitive to cyanide than the contractile process. Bradycardia caused by cyanide may be related in part to acetylcholine release since atropine (an acetylcholine blocker) can provide some protection against the toxic action of cyanide [30].

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13.2.3 LIVER AND KIDNEY Rabbits fed a mash containing 700 ppm inorganic cyanide for four weeks showed degenerative changes in liver and kidney [31]. Liver enzyme levels (e.g., alkaline phosphatase) were decreased with corresponding increases in serum levels. Also kidney alkaline phosphatase activity was decreased and serum urea and creatinine were increased reflecting impaired kidney function. Thus kidney and liver damage may occur with high chronic cyanide exposure. However these effects are not likely to be lifethreatening since kidneys have an excess of functional elements and liver tissue can be regenerated.

13.3 SYMPTOMS AND SIGNS CAUSED BY CYANIDE 13.3.1 ACUTE With lethal doses (200 mg KCN for an adult human) cyanide is rapidly fatal even by the oral route. Consciousness is lost in about 5 min; however, shallow, rapid respiration (due to carotid sinus chemoreceptor stimulation) may continue for several minutes following loss of consciousness. Deep stertorous breathing has also been observed in serious acute toxicity [2]. The heart continues to beat even after respiration stops, suggesting that neural tissue is more sensitive than the heart to cyanide. Foaming at the mouth and anoxic convulsions may occur as terminal events [2]. One would think that blockade of neural mechanisms for breathing is the key action of cyanide since death is due to asphyxia. Impaired neural function however results from poor circulation as well as a direct effect of cyanide on neural systems [1]. Thus maintenance of proper cardiovascular function is also important in treatment. Animal experiments provide insight into the human toxicology of cyanide. Laboratory mice given a sublethal dose of 5 mg/kg KCN subcutaneously become quiescent but resume normal locomotor activity after a few minutes [32]. They do not lose consciousness and can respond to a physical stimulus even during the quiescent period. These are important observations. A strong dose of 5 mg/kg has profound brain effects reflected by the overall decrease in motor activity, yet consciousness is maintained and the animals respond to prodding. Thus, with small increases in brain cyanide concentrations sensory and motor systems can still function normally, even after doses about 50% of the LD50 . Furthermore after a few minutes, animals resume spontaneous motor activity and behave normally. These acute effects of high doses are therefore transient in nature and probably independent of any cardiovascular actions.

13.3.2 AFTEREFFECTS OF ACUTE EXPOSURE Following unsuccessful suicide attempts with cyanide, patients may develop parkinsonian symptoms [33–37]. Several days to weeks after recovery from the acute effects, tremor, rigidity, and sustained abnormal muscle contractions appear. Treatment with L-DOPA (a drug also used for parkinsonism), shows variable improvement. Computerized tomography and magnetic resonance imaging show lesions in substantia nigra, cerebellum, and cerebral cortex. The location of lesions (substantia nigra, pars compacta) in idiopathic (without known cause) Parkinsons differs from that after cyanide intoxication (pars reticularis) [38]. More recently Jones et al. [39] reported that cyanide and dopamine interact to enhance cell death (apoptosis) of rat brain cells controlling motor movement, suggesting that cyanide may even contribute to idiopathic parkinsonism despite the fact that cyanide alone does not damage the pars compacta.

13.3.3 CHRONIC TOXICITY A variety of neurological diseases are proposed to result from chronic cyanide exposure [38,40]. Tropical ataxic neuropathy and “Konzo” are motoneuron diseases which occur in Africa among

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people who have cassava (edible rootstock) in their diets. When not properly prepared, the cassava can release cyanide. Simultaneous nutritional deficiencies may make the individual more sensitive to cyanide by decreasing cyanide metabolism. Chronic exposure of animals to cyanide can similarly cause brain lesions. Soto-Blanco and Gorniak [41] fed cyanide to female goats during lactation (1, 2, and 3 mg/kg/day for 90 days). Microscopic lesions were found in brain, thyroid, liver, and kidneys of both dams and kids. Soto-Blanco et al. [42] also fed KCN to rats (0.15, 0.3, and 0.6 mg/kg/day for 90 days). Lesions in ventral spinal cord and neuron loss in the hippocampus and cerebellum were noted. Thus animal experiments support the idea that chronic exposure to cyanide can cause neurodegenerative changes.

13.4 GENERATION AND METABOLISM OF CYANIDE IN BODY TISSUES Human blood contains about 3 µM cyanide and people who smoke have about 7 µM cyanide in their blood [43,44]. Consequently, most humans have about 0.5 to 1 mg of cyanide circulating in their bloodstream. Stelmaszynska and Zgliczynski [45] suggest that cyanide present in blood and excreted in urine and expired air may be produced, in part, by white blood cells, which generate cyanide in the process of destroying microorganisms, that is, phagocytosis. Cyanogenic bacteria in the gut are another potential source of blood cyanide [40,46]. Diet may influence blood cyanide since some foods naturally contain small amounts of cyanide (e.g., sweet potatoes and cabbage). Body tissues, for example, brain cells, can also produce cyanide [47–49]. The concentration of cyanide in the brain is about the same as in blood. Furthermore cyanide generated in the brain can influence neuronal activity [49]. Thus, small increases in cyanide concentrations in the brain can alter neuronal function. However the physiological purpose of brain cyanide is not known. One suggestion is that it regulates metabolism and may be a protective mechanism in animals (e.g., the toad) when availability of oxygen is limited [50]. Cyanide is a member of a small group of gaseous neuromodulators including nitric oxide, hydrogen sulfide, and carbon monoxide. These are di- or tri-atomic molecules that can be generated in the brain. They have good lipid and water solubility and can modulate activity in a group of neurons perhaps involving 1 to 2 million synapses [51]. Thus they contrast with classical neurotransmitters which operate on a more exacting, limited scale. In contrast to cyanide the neuromodulatory roles for nitric oxide [52], hydrogen sulfide [53], and carbon monoxide [54,55] are well established. Since cyanide (HCN) is a small molecule with good lipid and water solubility it is widely distributed in human tissues, it easily penetrates lipid membranes, and can act intracellularly. At physiological pH, over 98% of the molecule is in the form of HCN and only a small fraction occurs as cyanide anion (CN− ) (see Chapter 5). The major biological effects are due to undissociated HCN. It is likely that cyanide is formed in many body tissues, yet there are no reports in the literature of cyanide generation in any tissue except the brain and white blood cells. If cyanide does indeed have a physiological function, then mechanisms for regulating its levels should be present. Enzymes (rhodanese and 3-mercaptopyruvate sulfurtransferase) exist in mammalian systems to metabolize and inactivate cyanide by forming the less toxic thiocyanate [56]. They are most concentrated in liver and kidney but are also found in other tissues as well [57,58]. Rhodanese is a mitochondrial enzyme and is found in brain glial cells. The 3-mercaptopyruvate sulfurtransferase is located both in mitochondria and in cytoplasm as well as in blood [59]. These enzymes do not completely eliminate cyanide from body tissues, possibly because of limited availability of sulfur donors. (Other enzyme systems like cholinesterase reduce blood acetylcholine levels to nil.) Although cyanide concentration in whole blood may remain elevated for hours after a nonlethal exposure by inhalation [60], impaired enzymatic degradation of cyanide may increase susceptibility to cyanide toxicity [58].

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13.5 MECHANISMS OF CYANIDE TOXICITY 13.5.1 ACIDOSIS The immediate effect of acute high dose cyanide exposure involves direct effects on the brain and heart but at lower doses death is delayed and lactic acidosis becomes a significant factor in morbidity and mortality. Cyanide blocks oxidative metabolism and ATP generation, so glucose breakdown increases in the “glycolytic” pathway to compensate, and large amounts of lactic acid are produced in the process. Body pH decreases and many physiological processes are impaired. Hydrogen ion blocks calcium channels so heart and smooth muscle cells can not contract normally. Neither muscle cells nor neurons function well in an acid medium. Thus toxic mechanisms vary with dose and survival time after exposure to cyanide. If death is delayed acidosis becomes a significant factor. Some have even suggested that high plasma lactate is indicative of cyanide poisoning in patients and can be used to estimate the severity of the intoxication [61,62].

13.5.2 NEUROTRANSMITTER RELEASE Cyanide increases intracellular calcium and enhances release of various neurotransmitters [28,63,64]. This may explain some of the neurological symptoms such as convulsions, decreased motor activity, and tremors which are seen in cyanide intoxication [65,66]. Release of a mixture of inhibitory and excitatory neurotransmitters may explain many of the complex neurological symptoms of both acute and chronic cyanide toxicity.

13.5.3 ALTERATION OF CALCIUM BALANCE Calcium levels in neuronal cells increase in a few minutes after cyanide treatment [67]. The rapid sustained increase in free calcium caused by cyanide is due to both mobilization of intracellular and influx of extracellular calcium [68–70] (Figure 13.2). Cyanide (1 to 100 µM) activates an enzyme in nerve cell membrane (phospholipase C) which breaks down lipid in the membrane to generate inositol trisphosphate (IP3 ) [71]. IP3 mobilizes intracellular calcium. Influx of extracellular calcium occurs mainly by activation of certain receptors (“NMDA”) and NMDA receptor blockers are very effective in preventing cyanide-induced neuronal cell death [72,73]. Also calcium channel blockers can prevent the increase in total brain calcium and attenuate tremors caused by cyanide in mice [65]. Glutamate is the major excitatory neurotransmitter in the brain and is released by the action of cyanide. Glutamate release also occurs during brain anoxia (stroke) and triggers impaired function and neuronal death [74]. Thus the mechanism of glutamate release is critical to neuropathology. When metabolism is blocked as in cyanide poisoning or stroke, the high energy compound ATP diminishes and ion pumps (especially Na+ /K+ ATPase) are impaired. K+ accumulates extracellularly depolarizing the cells which reverses the Na+ gradient and causes the glutamate transporter to move glutamate to the outside of the cells. Activation of specific glutamate receptors (NMDA) by extracellular glutamate further depolarizes the cells to release more K+ and more glutamate as well. Blocking the anoxic depolarization current or glutamate receptors or the glutamate transporter prevents intracellular calcium accumulation and cell damage by anoxia [74]. These are key events in nerve damage caused by cyanide and by anoxia.

13.5.4 OXIDATIVE STRESS Antioxidants prevent cyanide-induced death of isolated neurons showing that reactive oxygen species (ROS) are important in the neurotoxic action of cyanide [75]. The oxidative stress is related to the rise in cytosolic Ca2+ since generation of intracellular oxidants (nitric oxide and other oxidative species) by cyanide is partly Ca2+ -dependent [75]. Some of the ROS is generated by the enzyme

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Ca2+

VSCC PLC

IP3

Ca2+ NMDA receptor Ca2+ Glu

FIGURE 13.2 Illustrating major toxic effects of calcium on neuronal calcium. Activation of phospholipase C to act on the membrane phosphatidyl inositol gives rise to IP3 which mobilizes intracellular calcium. Cyanide allows extracellular calcium to enter by way of voltage-sensitive channels (VSCC) and also though glutamate (Glu)-operated N-methyl-D-aspartate channels.

cyclooxygenase after cyanide treatment since aspirin (a cyclooxygenase blocker) decreases ROS levels and also prevents cell death [76]. It is interesting that taking an aspirin tablet may be beneficial in cyanide intoxication.

13.5.5 CYANIDE AND MITOCHONDRIA Mitochondria supply energy for cell function and also take up Ca2+ to buffer any excess in the cytoplasm. These processes are vital for normal function and for cell viability. Cyanide can block both these mechanisms by poisoning cytochrome oxidase (terminal enzyme in the electron transport chain) and limiting availability of ATP [46,77]. Accumulation of Ca2+ in mitochondria can be a serious toxic hazard. Glutamate-induced nerve damage results from Ca2+ loading and loss of mitochondrial function [77–79]. When used alone in low concentrations (<300 µM) cyanide does not completely block mitochondrial energy production, and some ATP remains to allow Ca2+ uptake into mitochondria. If opening of the mitochondrial permeability transition (PT) pore occurs during cyanide exposure, then more Ca2+ can enter. With excessive Ca2+ accumulation, mitochondria cease to function, ATP production falls, and cell death ensues. When the PT pore opens, the outer mitochondrial membrane becomes permeable and the protein cytochrome c escapes into the cytoplasm. Cytochrome c can initiate a series of reactions culminating in cell death. Blockade of mitochondrial PT using cyclosporin A can prevent cyanide-induced cytochrome c release and nerve cell death showing that the opening of the pore is a critical step in the toxic action of cyanide [39].

13.5.6 NERVE CELL DEATH PATHWAYS AND CYANIDE When mice were treated with 6 mg/kg KCN intraperitoneally twice daily for three days, two distinct lesions were found in brain sections [80]. In the suprarhinal cortex, cells died by a programmed cell death process (apoptosis), whereas cells in substantia nigra died by necrosis. Cultures of brain cells from these two areas taken from rat embryos also show a similar difference in response to cyanide

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treatment [81]. Thus the same concentration of cyanide has contrasting effects on neurons from different brain areas of rats and mice and presumably humans as well. Studies outlined below show that different proteins resident in these two cell types determine which death pathway is chosen when cyanide exposure occurs.

13.5.7 NEURONAL PROTEINS THAT INFLUENCE THE TOXIC EFFECTS OF CYANIDE “Bcl-2 family” proteins are important in mediating cyanide-induced apoptosis. Cyanide causes translocation of proteins (e.g., Bax) to mitochondria from the cell cytoplasm [82]. These proteins counteract the stabilizing effect of other Bcl-2 members which reside in the outer mitochondrial membrane (e.g., Bcl-2 itself and Bcl-XL ). This allows pore opening and escape of cytochrome c to the cytosol and initiation of programmed cell death. Thus ratios of these proteins, for example, Bax/Bcl-2 are critical in determining life and death of the cell after cyanide exposure. Uncoupling protein-2 (UCP-2) is expressed in the inner mitochondrial membrane in neuronal cells [83,84] and can influence the cell’s choice of a death pathway. Overexpression of UCP-2 leads to a rapid fall in mitochondrial membrane potential, a reduction in ATP, and death by necrosis [85]. UCP-2 overexpressing transgenic mice show a lower level of neuronal apoptosis following brain lesions [86]. Recent studies show that cyanide activates UCP-2 in the mitochondrial matrix [87]. It appears that large amounts of UCP-2 favor necrosis of a nerve cell in response to cyanide. In our laboratory, cultured primary mesencephalic brain cells from rat embryo, which are analogous to substantia nigral cells, show a more marked increase in UCP-2 in response to cyanide compared to cortical cells. These cells die of necrosis after cyanide treatment. So the protein UCP-2 appears to force substantia nigral cells into a necrotic pathway. Another neuronal mitochondrial protein which appears to influence cell death is BNIP3. This protein is a member of the Bcl-2 family and can influence both apoptosis and necrosis [88–91]. BNIP3 can induce necrosis by interaction with Bcl-2/Bcl-XL to neutralize their antiapoptotic actions [88,92]. The mode of cell death activated by BNIP3 may depend on its expression level and the degree of interaction with Bcl-2 proteins. Cyanide increases BNIP3 in both cortical and mesencephalic cells and the protein acts cooperatively with UCP-2 to induce necrosis [91]. In our lab cyanide increases BNIP3 levels in both cortical and mesencephalic cells. Also when the RNA interference probe for BNIP3 is used in these cells, there is a decrease of BNIP3 and an overall decrease in death in both cortical and mesencephalic cells (unpublished data). Thus BNIP3 like UCP-2 has a great influence on death processes in brain cells. These two proteins may coordinate to determine cell survival and pathways of cell destruction after cyanide treatment. New insight into mechanisms of cyanide-induced cell death may lead to improved antidotes for this powerful agent.

13.6 CYANIDE ANTIDOTES Cyanide has caused many deaths. Poisonings still occur commonly, primarily in relation to smoke inhalation, so there is a continuing need for cyanide antidotes. Cyanide response kits available in the United States typically contain three antidotes: amyl nitrite, sodium nitrite, and sodium thiosulfate. Amyl nitrite is a volatile liquid; the glass vial is crushed in gauze to allow inhalation by the comatose patient. The drug can give rise to nitric oxide which appears to antagonize NMDA receptor activation and can reverse the effect of cyanide on the cytochrome oxidase enzyme [93,94]. Sodium nitrite is given intravenously to increase levels of methemoglobin, a cyanide scavenger. Thiosulfate supplies sulfur for conversion of cyanide to thiocyanate. The nitrite/thiosulfate regimen can protect against six LD50 s of cyanide in animals [95]. Sodium nitrite may be dangerous as a cyanide antidote since it can cause excessive methemoglobin formation with fatalities [96]. Also it must be given intravenously and thus is not suitable for out-of-hospital use. The cyanide kit is expensive ($100) and not well stocked by many hospitals [97].

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Some believe that we are poorly prepared for dealing with cyanide intoxication on a large scale which could occur in terrorist attacks [97]. Alpha ketoglutarate plays an important role biochemically in amino acid transamination and has been recognized as an effective cyanide scavenger by cyanohydrin formation [98] but has not been employed clinically. Hydroxocobalamin (Vitamin B12 ) is used in Europe as a cyanide antidote. The cobalt chelates with cyanide to neutralize its effect. It is nontoxic and safe at high doses. Unfortunately it is not available in the United States. For immediate treatment of cyanide intoxication, Borron and Baud [99] recommend use of 100% oxygen, assisted ventilation, decontamination, correction of the acidosis, and blood pressure support. They indicate that hydroxocobalamin is an attractive antidote due to its rapid cyanide binding and lack of serious side effects. Also they point out that sodium thiosulfate acts more slowly than other antidotes and is more appropriate for subacute poisoning or supportive treatment. Improvement of blood pressure and blood pH reflect effectiveness of antidotal therapy but the final outcome depends on the extent of brain injury. There is a critical need to develop a new cyanide antidote that could be used for medical treatment of cyanide-exposed individuals. The available antidotes are either scavengers (binders) of cyanide or increase the rate of elimination of cyanide through biotransformation. A direct antagonist of cyanide biochemical action has not been developed. The antidotes presently available have a low efficacy and many disadvantages. However, significant advances have been made in recent years in understanding the toxicodynamics of cyanide, which provides for a basis to design a new class of antagonists. The primary toxicological actions of cyanide are in the mitochondrial matrix, so targeting of compounds to the mitochondria to reverse the toxicity by inhibiting the activation of UCP-2 may prove to be a novel, effective means of directly reversing the biochemical action of cyanide. Cyanide is a powerful poison and can cause profound effects on vital organs. Timely, effective use of antidotes is important to avoid a fatal outcome and also to prevent neurodegenerative changes.

13.7 SUMMARY AND CONCLUSIONS • Cyanide, an old, well-known poison is in widespread industrial use, is present in the environment, is generated by combustion of man-made materials, and is a potential terrorist threat. • Hydrogen cyanide gas (HCNg ), free cyanide in aqueous solution (HCNaq , and CN− which can be converted to HCNaq ), and simple cyanide salts (e.g., NaCN, KCN) are the primary toxic forms of cyanide. • The brain and the heart are important sites of action of cyanide in humans. • Cyanide is lethal in high doses (200 mg KCN in an adult human) but lower doses can damage sensitive brain cells. • Cyanide interferes with energy production and causes oxidative stress by acting on mitochondria in brain cells. • Probably related to neurotransmitter release, cyanide can produce irregular heart beats, slow heart rate, and cause heart failure. • Proteins (“UCP-2” and “BNIP3”) in cell mitochondria may mediate some toxic actions of cyanide. • Several cyanide antidotes are available, but none is ideal. Methemoglobin (combines with cyanide) formers are effective but too much methemoglobin is toxic. Sulfur donors (thiosulfate) promote conversion of cyanide to thiocyanate but must be given intravenously. Hydroxocobalamin combines with and neutralizes cyanide but is not available in the United States.

Acknowledgement The cited research by the authors was supported in part by NIH grant ESO 4140.

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14 Aquatic Toxicity of Cyanide Robert W. Gensemer, David K. DeForest, Angela J. Stenhouse, Cortney J. Higgins, and Rick D. Cardwell CONTENTS 14.1 14.2

Mechanisms of Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cyanide Bioavailability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.2.1 Evidence Supporting Toxicity of Free versus Complex Cyanides . . . . . . . . . . . . . . 14.2.2 Other Environmental Factors that Influence Bioavailability . . . . . . . . . . . . . . . . . . . . 14.3 Toxicity to Freshwater Biota. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.1 Acute Lethality to Freshwater Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.2 Acute Toxicity to Freshwater Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.3 Acute Sublethal Effects on Freshwater Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.4 Chronic Toxicity to Freshwater Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.5 Chronic Toxicity to Freshwater Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3.6 Freshwater Algae and Vascular Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4 Toxicity to Marine Biota . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.1 Acute Toxicity to Marine Fish. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.2 Acute Toxicity to Marine Invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.3 Chronic Toxicity to Marine Fish and Invertebrates. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.4 Marine Algae and Vascular Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5 Regulatory Criteria for Protection of Aquatic Life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.1 Basis of Current AWQC . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.2 Comparison of Current AWQC to Updated Toxicity Literature . . . . . . . . . . . . . . . . 14.5.3 Criteria Implementation as a Function of Cyanide Chemical Form . . . . . . . . . . . . 14.6 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

252 252 253 254 254 270 270 271 272 273 273 273 273 274 274 275 275 276 276 277 278 279 279

Concerns over the toxicity of cyanide to aquatic organisms have resulted from the use and discharge of cyanide compounds from sewage treatment plants, industrial processes, and from extraction of gold or silver from natural ores [Chapter 3; 1]. Free hydrogen cyanide (HCN) concentrations can occur in surface waters from 0.005 to several mg HCN/l [1], and thus, may reach concentrations that are sublethal or lethal to both freshwater and marine organisms under some circumstances. To help quantify ecological risks to aquatic organisms from exposure to aqueous cyanides, aquatic toxicity data have been compiled to develop ambient water quality criteria for the protection of aquatic life in the United States [2]. National ambient water quality criteria (AWQC) set maximum threshold concentrations of inorganic and organic contaminants, for both freshwater and marine environments [3]. These criteria are derived from empirical toxicity data, and are designed to protect all but the 251

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most sensitive 5% of species in an aquatic community, although criteria may be lowered further to ensure protection of ecologically, economically, or recreationally important species. Recently, concerns have arisen that the AWQC for cyanide have been problematic to implement, because they may not accurately reflect either the toxic forms or bioavailable1 concentrations of cyanide in water, sediments, and the tissues of aquatic organisms. These cyanide criteria typically have been implemented on the basis of total cyanide concentrations, rather than free cyanide concentrations that formed the basis of the criteria calculations upon which the criteria were derived. Therefore, this chapter reviews existing data and presents new knowledge on cyanide toxicity, cyanide chemical speciation and its measurement, and the relative toxicity of bioavailable cyanide species. Our primary goal is to evaluate current criteria concentrations relative to new studies conducted since the 1984 AWQC [2], and to evaluate what chemical measurements used for compliance with AWQC would best reflect a current understanding of cyanide bioavailability to aquatic organisms.

14.1 MECHANISMS OF TOXICITY Hydrogen cyanide, HCN, is a highly toxic and highly soluble chemical that is readily taken up by aquatic organisms via contact with skin and mucous membranes (e.g., gills), from where it is readily absorbed into the bloodstream [1]. The transport of HCN from the water through biological membranes into the blood stream is facilitated by the small size of the HCN molecule, and its lack of an ionic charge [5]. Upon uptake by an organism, HCN acts as a powerful inhibitor of the haemoprotein, cytochrome oxidase [6]. Cyanide binds preferentially to the iron porphyrins present in cytochrome oxidase, which in turn withholds the oxygen in the blood from tissues, and prevents mitochondrial haemoprotein from accepting electrons through the electron transport system. Above toxic thresholds, cyanide will cause death by histotoxic apoxia, while sublethal doses can result in an organism falling into an anesthetized state [6]. This anesthetized state, induced by sublethal concentrations of cyanide, has resulted in its use, often illegally, to collect tropical marine fish for aquariums and general markets (Chapter 11). When fish are collected in this manner, they are exposed to very high concentrations of cyanide (>5000 µg/l) for a short period of time. Some fish die upon exposure, while others are “anesthetized,” and can recover when placed in cyanide-free water. The primary mechanism by which organisms detoxify cyanide is through its conversion to thiocyanate (SCN− ). As thermodynamics favor the formation of thiocyanate, there is limited reverse conversion to cyanide [7]. Thiocyanate is readily excreted from an organism through urine; however, the kidney tubules are also very efficient at reabsorbing SCN− [8]. This results in slow and irregular SCN− excretion. While less toxic than cyanide, SCN− also imparts toxic effects, due to its antithyroidal properties [7]. While SCN− can have histological effects on the thyroid, and reproductive effects on Pimephales promelas (fathead minnows) at concentrations greater than 1000 and 7000 µg/l, respectively, the amount of cyanide required to bring about SCN− levels this high would be fatal [7].

14.2 CYANIDE BIOAVAILABILITY Given that the primary mode of toxic action for cyanide involves the transport of the HCN molecule into an organism’s bloodstream, it is reasonable to assume that concentrations of simple ionic forms of cyanide would best predict toxicity to aquatic organisms. While cyanide exists in a variety of metallocyanide or organic complexes in the aquatic environment (Chapter 2), the toxicity of these complexes is largely a function of their dissociation to free cyanide (i.e., HCN or CN− ). This section 1 The

term “bioavailability” is defined in this context as the degree to which a chemical can be taken up by an organism, subsequently interacting with a biologically important site of action [4].

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reviews the evidence supporting this assumption, and discusses other factors that might influence cyanide speciation or bioavailability to aquatic organisms.

14.2.1 EVIDENCE SUPPORTING TOXICITY OF FREE VERSUS COMPLEX CYANIDES Complex metallocyanide ions in solution can be dissociated or decomposed to release free cyanide ions that, in turn, form HCN through hydrolytic reactions in water [9]. Measuring only the concentration of molecular HCN will thus provide indirect evidence of acute and chronic toxicity of solutions containing complex metallocyanide ions, but will not be entirely conclusive. This is because the concentration of HCN in solution can change due to exposure to natural light, changes in pH or hardness, or because of increased dilution of the complexes [9,10]. Some metallo-cyanide complexes are unstable, and will dissociate quickly in dilute solutions, such as zinc cyanide or cadmium cyanide. The production of toxic metal ions and cyanide that dissociate to their free ionic states is the reason zinc cyanide and cadmium cyanide complexes have been shown to be exceedingly toxic to aquatic life [10]. Additionally, the Ag(CN)2− ion is, on its own, a fairly toxic compound, and can be predominant when its dissociation is very slight [9]. Nickel and copper readily bind with free cyanide, to form the relatively stable and nontoxic nickel–or copper–cyanide ion [10]. The nickel–cyanide complex is not significantly decomposed by direct sunlight, but will dissociate under specific pH conditions [10]. If the solution is alkaline (pH ≥ 8), there will be nominal dissociation of the nickel cyanide ion; as the pH decreases below 8, however, the dissociation of the ion increases, as does toxicity. As the pH approaches 6.5, the toxicity of cyanide combined with nickel approaches that of free cyanide [10]. The copper–cyanide complex is also quite stable, able to resist photodegradation, and does not readily dissociate even at a pH of 7.0. This complex completely dissociates only in extremely soft water that is continuously aerated [10]. The iron–cyanide complexes are probably the least toxic to aquatic life. Ferrocyanide ions are also the most stable metallocyanide complexes; however, they are subject to photolysis by UV light [9; see also Chapter 5]. Studies conducted by Doudoroff [10], Doudoroff et al. [9], and Pablo et al. [11,12] have also shown that HCN dissociated from the metallocyanides is the source of toxicity observed in aquatic organisms exposed to metallocyanide complexes, rather than total cyanide. Doudoroff et al. [9] evaluated the median resistance time of young bluegill sunfish (Lepomis macrochirus) to various concentrations of NaCN, NaCN + NiSO4 , NaCN + CuSO4 , NaCN + ZnSO4 , NaCN + CdSO4 , and NaCN + AgNO3 . This study measured total and free cyanide in the test solutions. They found that the total and free cyanide concentrations were nearly equal, in treatments receiving only NaCN and low concentrations of NaCN + ZnSO4 and NaCN + CdSO4 . The dilutions containing NaCN + NiSO4 , NaCN + CuSO4 , and NaCN + AgNO3 had much lower concentrations of HCN than total cyanide. The toxic effects imparted on the bluegill by these metallocyanides were, however, correlated to the concentration of HCN measured in the test solutions. The sole exception was AgNO3 , which imparted toxicity in addition to that of the dissociated cyanide. Pablo et al. [11] compared the toxicity of NaCN, K3 Fe(CN)6 , and K4 Fe(CN)6 on the embryonic development of the doughboy scallop (Chlamys asperrimus). Again, the concentration of total cyanide and HCN was measured at the initiation and the end of the test. The order of toxicity based on total cyanides to the scallop larvae was NaCN > K3 Fe(CN)6 > K4 Fe(CN)6 , with LC50s (median lethal concentrations) of 28.6, 128, and 160 µg/l, respectively. Corresponding LOECs (lowest observable effects concentrations) were 10, 45, and 60 µg/l, respectively. If the LC50s are compared using free cyanide concentrations they are 8.4, 22.1, and 24.3 µg/l for NaCN, K3 Fe(CN)6 , and K4 Fe(CN)6 , respectively. When LC50s were plotted against measured 48-h average free cyanide concentrations, the toxicity of the iron–cyanide complexes correlated well with the toxicity of NaCN. This suggests that the toxicity of the complexes was due only to the free cyanide component.

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Pablo et al. [12] conducted a similar study comparing the toxicity effects of NaCN, K3 Fe(CN)6 , and K4 Fe(CN)6 on the survival of the banana prawn (Penaeus monodon). Again, there was a good correlation between the toxicities of K3 Fe(CN)6 , K4 Fe(CN)6 , and NaCN to P. monodon, based on measured free cyanides. Though these metallocyanides proved to be slightly more toxic than predicted, the results again indicated that the primary source of toxicity observed in all treatments was free cyanide. Although the complex anions [FeIII (CN)6 ]3− and [FeII (CN)6 ]4− imparted insignificant toxicity, this slight toxicity could have been due to their dissociation to ferric or ferrous CN− ions (and, consequently, HCN) when HCN penetrates the cells, or when the organisms absorb iron. Collectively, the Doudoroff and Pablo studies support the hypothesis that cyanide bound to metals such as iron, copper, or nickel is much less toxic to both freshwater and marine organisms, than when present individually in its free ionic form. Additionally, toxicity observed in tests using these metallocyanide complexes was not, with the exception of silver–cyanide complexes, due to these complex ions or total cyanide, but rather, was due to the concentration of free cyanide (HCN), and to the dissociated metallic ions.

14.2.2 OTHER ENVIRONMENTAL FACTORS THAT INFLUENCE BIOAVAILABILITY Cyanide toxicity to aquatic organisms has been reported to vary as a function of water pH, temperature, and dissolved oxygen concentrations [1], but these relationships were not considered to be strong enough to derive AWQC as a mathematical function of these factors [2]. While pH can have a strong influence on the dissociation of inorganic cyanide complexes (Section 14.2.1; Chapter 5), the toxicity of the free cyanides that result from this dissociation does not vary significantly within a pH range of 6.8 to 8.3. Cyanide can be more toxic to freshwater fish at low dissolved-oxygen concentrations [1], but so long as toxicity tests are conducted properly (i.e., sufficient dissolved-oxygen concentrations are maintained), no adjustments to toxicity criteria as a function of dissolved oxygen should be necessary. Temperature has been shown to significantly affect cyanide toxicity in teleost fishes, but its influence differs at sublethal (toxicity greatest at low temperatures) versus rapidly lethal (toxicity greatest at elevated temperatures) concentrations of cyanide [1,2]. No studies conducted since publication of the U.S. Environmental Protection Agency (USEPA’s) 1984 AWQC have further evaluated the influence of these or other water quality characteristics on cyanide toxicity to aquatic organisms.

14.3 TOXICITY TO FRESHWATER BIOTA As discussed in Section 14.2, cyanide toxicity to aquatic life is predominantly a function of free, rather than total, cyanide concentrations. Cyanide tends to be rapidly metabolized and excreted (Section 14.1), and so, does not tend to bioaccumulate or biomagnify (Chapter 17). Accordingly, exposure of aquatic life to cyanide via the sediment or dietary pathway may be relatively unimportant, compared to water-based exposure. Furthermore, cyanide exposure via these pathways has received little attention in the scientific literature, so our review below considers cyanide toxicity to freshwater biota via aqueous exposure only. Our primary source of aqueous toxicity data was the USEPA’s AWQC for cyanide [2], which was then supplemented with more recent studies from the primary scientific literature. All acute and chronic toxicity data considered acceptable for derivation of AWQC [3] were tabulated separately (Tables 14.1–14.5) from “other” data (Table 14.2) that were generated using experimental designs or test endpoints not considered sufficiently reliable or relevant for this purpose according to USEPA guidance. Our only significant deviation from USEPA guidance in compiling these tables was reliance upon data for all species, regardless of their geographic distribution (i.e., not necessarily residents of North America only). Tables 14.1 to 14.5 include data from the 1984 AWQC, as well as newer data when available, except that for purposes of brevity, Table 14.2 includes only data from studies

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Cichlidae Tilapia mossambica (Mozambique tilapia)

Micropterus salmoides (largemouth bass) Pomoxis nigromaculatus (black crappie)

Fish Centrarchidae Lepomis macrochirus (bluegill sunfish)

Species

NaCN NaCN NaCN

NaCN

NR

19 g 19 g 19 g

KCN KCN KCN KCN KCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN

Cyanide form

Juvenile 3.88 cm, 0.96 g 6.09 cm, 2.80 g 14.24 cm, 54.26 g Juvenile Juvenile Fry Fry Fry Fry Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile

Age/size

96-h 96-h 96-h

96-h

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

Exposure duration

U U U

Note 1

U U U U Note 3 M-NR Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1

Analytical methoda

LC50 LC50 LC50

LC50

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

Endpoint

TABLE 14.1 Acute Toxicity of Cyanide to Freshwater Aquatic Animals (Acceptable According to USEPA [3])

Mortality Mortality Mortality

Mortality

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Effect

1067 1045 194

102

180 220 180 230 160 150 364 232 279 273 81 85.7 74 100 107 99 113 121 126 102

Concentration (µg CN/l)

[52] [52] [52]

[51]

[47–49] [47–49] [47–49] [47–49] [47–49] [50] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51]

Reference

Aquatic Toxicity of Cyanide 255

Lepidocephalichithys guntea (peppered loach) Pimephales promelas (fathead minnow)

Labeo rohita (rohu)

Labeo calbasu (orange-fin labeo)

Labeo bata (bata)

Cirrhinus mrigala (mrigal)

Cyprinidae Carassius auratus (goldfish) Catla Catla (catla)

Species

TABLE 14.1 Continued

NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN

Juvenile Juvenile Juvenile Juvenile 02.–05.g Juvenile Juvenile Fry Fry Fry Fry

Cyanide form

Juvenile 27 g 27 g 27 g 18 g 18 g 18 g 21 g 21 g 21 g 17 g 17 g 17 g 26 g 26 g 26 g 5.2–5.5 cm

Age/size

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

Exposure duration

Note 1 Note 1 Note 1 Note 5 U M-NR M-NR Note 1 Note 1 Note 1 Note 1

Note 4 U U U U U U U U U U U U U U U U

Analytical methoda

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

Endpoint

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Effect

120 113 128 230 170 350 230 120 98.7 81.8 110

318 934 918 294 807 838 196 998 1045 249 1051 1029 215 1035 1045 199 175.2

Concentration (µg CN/l)

[55] [55] [55] [10] [17] [50] [50] [51] [51] [51] [51]

[53] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [52] [54]

Reference

256 Cyanide in Water and Soil

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Salmonidae Oncorhynchus mykiss (rainbow trout)

Poeciliidae Poecilia reticulata (guppy)

Percidae Perca flavescens (yellow perch)

Fry Juvenile Juvenile Juvenile Juvenile

Adult

Embryo Fry Fry Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile

Fry Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile

NaCN NR NR NR NaCN

KCN

NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN

NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN

96-h 96-h 96-h 96-h 96-h

96-h

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

U Note 7 Note 7 Note 7 U

Note 6

Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1

Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1

LC50 LC50 LC50 LC50 LC50

LC50

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

Mortality Mortality Mortality Mortality Mortality

Mortality

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

90 27 40 65 46.3

147

281 288 330 88.9 93 74.7 94.7 101 107

116 119 126 81.5 124 137 131 105 119 131 122 161 188 175 163 169

[57] [58] [58] [58] [59]

[56]

[51] [51] [51] [51] [51] [51] [51] [51] [51]

[51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51]

Aquatic Toxicity of Cyanide 257

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Salmo salar (Atlantic salmon) Salvelinus fontinalis (brook trout)

Species

TABLE 14.1 Continued Cyanide form NaCN NaCN NaCN NaCN KCN KCN KCN KCN NaCN KCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN NaCN

Age/size Juvenile Juvenile Juvenile Juvenile Fingerling Fingerling Fingerling Juvenile Juvenile Juvenile Adult Sac fry Sac fry Sac fry Sac fry Swim-up fry Swim-up fry Swim-up fry Swim-up fry Swim-up fry Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

Exposure duration U U U U Note 7 Note 7 Note 7 U Note 1 Note 8 Note 4 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1 Note 1

Analytical methoda LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

Endpoint Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Effect 52.1 54.1 62.1 74.8 53.3 53.3 41.7 97 57 90 156 105 342 507 252 84 54.4 86.5 104 90.3 73.5 83 75 86.4 91.9 99 96.7 112 52 60.2

Concentration (µg CN/l)

[59] [59] [59] [59] [46] [46] [46] [60] [51] [61] [53] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51] [51]

Reference

258 Cyanide in Water and Soil

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Daphnia spp. (cladoceran)

Daphnia pulex (cladoceran)

Daphniidae Daphnia magna (cladoceran)

Cyclopidae Cyclops viridis (copepod)

NR NR NR

NR NR 1st and 2nd instar NR >=6-d

NR NR NR

NR NR NR

3rd or 4th instar

Chironomidae Tanytarsus dissimilis (midge)

Corixidae Corixa sp. (water boatman)

NR 0.012 g

NR NR NR

Asellidae Asellus communis (isopod) Asellus intermedius (isopod)

Invertebrate Ambularidae Pila globosa (snail)

Juvenile Juvenile Juvenile Juvenile

NaCN NaCN NaCN

NaCN KCN NaCN KCN KCN

NaCN NaCN NaCN

NaCN NaCN NaCN

NaCN

NaCN NaCN

NaCN NaCN NaCN

NaCN NaCN NaCN NaCN

96-h 96-h 96-h

48-h 96-h 96-h 48-h 48-h

96-h 96-h 96-h

96-h 96-h 96-h

48-h

96-h 96-h

96-h 96-h 96-h

96-h 96-h 96-h 96-h

U U U

U U U M-NR U

U U U

U U U

Note 2

Note 1 U

U U U

Note 1 Note 1 Note 1 Note 1

LC50 LC50 LC50

LETC LC50 LC50 LC50 LC50

LC50 LC50 LC50

LC50 LC50 LC50

LC50

LC50 LC50

LC50 LC50 LC50

LC50 LC50 LC50 LC50

Mortality Mortality Mortality

Immobilization Mortality Mortality Mortality Mortality

Mortality Mortality Mortality

Mortality Mortality Mortality

Mortality

Mortality Mortality

Mortality Mortality Mortality

Mortality Mortality Mortality Mortality

80 169 172

1800 159.8 90 110 83

78 167 169

247 251 252

2490

2326 1700

1571 1539 891

66.8 71.4 97 143

[52] [52] [52]

[64] [65] [17] [16] [15]

[52] [52] [52]

[52] [52] [52]

[63]

[62] [17]

[52] [52] [52]

[51] [51] [51] [51]

Aquatic Toxicity of Cyanide 259

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Physidae Physa heterostropha (snail)

Ranatra sp. (water scorpion)

Nepidae Nepa sp. (water scorpion)

NR

NR NR NR NR NR NR

NR NR NR

0.006 g

Lumbriculidae Lumbriculus variegatus (oligochaete)

Lymnaeidae Lymnaea luteola (snail)

0.007 g NR

NR NR NR

NR NR NR

Age/size

Gammaridae Gammarus fasciatus (amphipod) Gammarus pseudolimnaeus (amphipod)

Dytiscidae Dytiscus sp. (aquatic beetle)

Diaptomodae Diaptomus sp. (copepod)

Species

TABLE 14.1 Continued

KCN

NaCN NaCN NaCN NaCN NaCN NaCN

NaCN NaCN NaCN

NaCN

NaCN NaCN

NaCN NaCN NaCN

NaCN NaCN NaCN

Cyanide form

96-h

96-h 96-h 96-h 96-h 96-h 96-h

96-h 96-h 96-h

96-h

96-h 96-h

96-h 96-h 96-h

96-h 96-h 96-h

Exposure duration

U

U U U U U U

U U U

U

U Note 1

U U U

U U U

Analytical methoda

LC50

LC50 LC50 LC50 LC50 LC50 LC50

LC50 LC50 LC50

LC50

LC50 LC50

LC50 LC50 LC50

LC50 LC50 LC50

Endpoint

Mortality

Mortality Mortality Mortality Mortality Mortality Mortality

Mortality Mortality Mortality

Mortality

Mortality Mortality

Mortality Mortality Mortality

Mortality Mortality Mortality

Effect

432

241 289 293 228 229 231

1316 1342 1315

11,000

900 167

246 249 259

82 166 172

Concentration (µg CN/l)

[47,49]

[52] [52] [52] [52] [52] [52]

[52] [52] [52]

[17]

[17] [62]

[52] [52] [52]

[52] [52] [52]

Reference

260 Cyanide in Water and Soil

NR NR NR

22 mm, 0.39 g

0.180 g NR NR NR

0.006 g

NaCN NaCN NaCN

NaCN

NaCN NaCN NaCN NaCN

NaCN

96-h 96-h 96-h

96-h

96-h 96-h 96-h 96-h

96-h

U U U

NR

U U U U

U

LC50 LC50 LC50

EC50

LC50 LC50 LC50 LC50

LC50

Mortality Mortality Mortality

Ability to detach from substrate

Mortality Mortality Mortality Mortality

Mortality

Note 1: Note 2: Note 3: Note 4: Note 5: Note 6: Note 7: Note 8: Note 9: Note 10: Note 11: Note 12: Note 13: Note 14: Note 15: Note 16: Note 17:

Free CN by Epstein colorimetric method [45]. Colorimetric method [67]. Tartaric acid distillation, followed by colorimetric method [68]. Pyridine-pyrazalone method [69]. Initial concentration verified based on titration with silver nitrate, using rhodamine indicator. Light spectrophotometry (Way, J.L., personal communication). Lambert et al. [70]. Mertens [71]. Pyridine-barbituric acid method with pH and temperature correction [67,72] . Flame spectrophotometry (after filtering). Weak acid dissociable method 4500-CN method I [73]. APHA [74]. Pyridine-pyrazolone spectrophotometric method [75]. Pyridine-barbituric acid method [74]. Spectrophotometric analysis sensitive to 20 µg/l in sea water [76]. Cyanide electrode. Colorimetric [77].

a Analytical notes: U = unmeasured, M-NR = measured, but not reported, NR = not reported. Detailed numeric notes are as follows:

Viviparidae Viviparus bengalensis (snail)

Pteronarcyidae Pteronarcys dorsata (stonefly)

Planorbidae Helisoma trivolvis (snail) Planorbis exustrus (snail)

Planariidae Dugesia tigrina (planarian)

1539 1576 1531

426

50,000 1581 1540 870

2100

[52] [52] [52]

[66]

[17] [52] [52] [52]

[17]

Aquatic Toxicity of Cyanide 261

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 261 — #11

HCN HCN

Adult Adult

NaCN

KCN KCN

NR

Pomacentridae Dascyllus aruanus (damselfish)

KCN

NaCN

egg and sperm Young

35–49 g

Percidae Perca fluviatilis (European perch)

Salmonidae Oncorhynchus mykiss (rainbow trout)

205 g

Mugilidae Mugil auratus (golden mullet)

KCN

KCN

44 mm fork length

9–10 cm

HCN

NR

Cyanide form

<12-h

NR

Age/size

Mormyridae Gnathonemus petersi (elephantnose fish)

Cyprinidae Pimephales promelas (fathead minnow) Rutilus rutilus (intersex roach)

Fish Centrarchidae Lepomis macrochirus (bluegill sunfish)

Species

TABLE 14.2 Other Cyanide Toxicity Data for Aquatic Organisms

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 262 — #12 7-d 7-d

10-d 42-d

120-sec

6-d

24-h

6-h

7-d

96-h

96-h

Exposure duration

Note 7 Note 7

Note 10 Note 12

U

Note 17

U

U

Note 17

Note 9

U

Analytical methoda

NOEC LOEC

LOEC NOEC

LOEC

LC50

NOEC

LOEC

LC50

LOEC

LC100

Endpoint

Reproduction Susceptibility to Saprolegnia parasitica Hepatosomatic index Serum calcium levels reduced

Mortality

Mortality

Plasma parameters

Gill damage

Mortality

Growth

Mortality

Effect

29 10

0 21

25,000

98

159

20

108

58

1000

Concentration (µg CN/l)

[83] [83]

[24] [82]

[6]

[80]

[81]

[18]

[80]

[79]

[78]

Reference

262 Cyanide in Water and Soil

Salmo salar Ouananiche (Landlocked Atlantic salmon)

HCN HCN HCN

2–3 yrs 2–3 yrs 2–3 yrs

HCN

HCN

2–3 yrs

Adult

HCN

2 yr

HCN

HCN HCN HCN HCN

Adult Adult Adult 2 yr

Adult

HCN

168.5 g

HCN HCN

HCN KCN KCN KCN KCN KCN KCN KCN

NR Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile Juvenile

Adult Adult

HCN

Adult

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 263 — #13 12-d

12-d

12-d 12-d

12-d 12-d 12-d

12-d

12-d

12-d 12-d 12-d 12-d

20-d

96-h 16-wk 16-wk 16-wk 16-wk 16-wk 21-d 21-d

7-d

Note 7

Note 7

Note 7 Note 7

Note 7 Note 7 Note 7

Note 7

Note 7

Note 7 Note 7 Note 7 Note 7

U

Note 13 Note 14 Note 14 Note 14 Note 14 Note 14 Note 16 Note 16

Note 7

LOEC (only conc. tested) No sign. effect (only conc. tested) No sign. effect (only conc. tested) No sign. effect (only conc. tested)

LOEC (only conc. tested) LOEC (only conc. tested) LOEC (only conc. tested)

No sign, effect (only conc. tested) LOEC (only conc. tested)

LOEC (only conc. tested) LOEC (only conc. tested) LOEC (only conc. tested) LOEC

LOEC

NOEC NOEC NOEC NOEC NOEC NOEC LOEC NOEC

NOEC

Gonadosomatic index

Hepatosomatic index

Plasma vitellogenin Liver vitellogenin

Diameters of vitellogenic oocytes Dopamine levels Norepinephrine levels Number of spermatogonial cysts

Serum phosphoprotein phosphorus Liver effects Growth Hepatosomatic index Splenosomatic index Hematocrit Plasma thyroid levels Increased plasma cortisol Blood cholesterol and sodium Liver cytochrome oxidase activity Plasma vitellogenin Gonadosomatic index Hepatosomatic index Number of spermatogonial cysts Gonadosomatic index

5

5

5 5

10 10 10

10

10

10 10 10 10

10

13 25 25 25 25 25 21 21

29

[27]

[27]

[27] [27]

[26] [26] [26]

[26]

[25]

[25] [25] [25] [25]

[85]

[84] [7] [7] [7] [7] [7] [20] [20]

[83]

Aquatic Toxicity of Cyanide 263

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 264 — #14

Haliotidae Haliotis varia (abalone)

Gammaridae Gammarus pulex (amphipod)

Faviidae Plesiastrea versipora (coral)

Daphnia pulex (cladoceran)

Daphniidae Daphnia magna (cladoceran)

Cambaridae Orconectes rusticus (crayfish)

Brachionidae Brachionus calyciflorus (rotifer)

Invertebrate Arcidae Scapharca inaequivalvis (clam)

Species

TABLE 14.2 Continued

33–50 mm shell 33–50 mm shell

1–1.5 cm 1–1.5 cm 1–1.5 cm 1–1.5 cm 1–1.5 cm

Mature Mature Mature

NR <24-h <24-h

NR

NR

4 cm shell length

Age/size

KCN NaCN

KCN KCN KCN KCN KCN

NaCN NaCN NaCN

KCN KCN KCN

NR

KCN

NaCN

Cyanide form

96-h 96-h

11.6–19.4-h 7–11.6-h 4.9–7.8-h 4.1–7.9-h 2.5–3.4-h

3-h 3-h 3-h

24-h 24-h 24-h

9610h

24-h

11.2-d

Exposure duration

U U

U U U U U

U U U

U U U

U

U

U

Analytical methoda

LOEC LOEC

LT50 LT50 LT50 LT50 LT50

NOEC-LOEC LC100 NOEC-LOEC

EC50 EC50 EC50

LC100

LC50

LT50

Endpoint

Mortality Mortality

Mortality Mortality Mortality Mortality Mortality

Fluorescence Mortality No. of Zooanthellae

Immobilization Immobilization Immobilization

Mortality

Mortality

Mortality

Effect

400 1062

1199 2997 5993 11,987 29,967

26–260 260,136 26–260

401 243 223

5000

62,446

26,000

Concentration (µg CN/l)

[90] [90]

[89] [89] [89] [89] [89]

[88] [88] [88]

[86] [87] [87]

[78]

[86]

[39]

Reference

264 Cyanide in Water and Soil

KCN

NaCN

Mature

NR

NaCN

Mature

<6-h <6-h

budding budding

48-h

10-sec

30-min

120-h 120-h

144-h 144-h

U

U

U

Chronic Chronic

Chronic Chronic

LC50

LOEC

LOEC

LOEC NOEC

NOEC NOEC

Mortality

Photosynthesis of symbiotic algae

Zooxanthellae density

Reproduction Mortality

Mortality Reproduction

Note 1: Note 2: Note 3: Note 4: Note 5: Note 6: Note 7: Note 8: Note 9: Note 10: Note 11: Note 12: Note 13: Note 14: Note 15: Note 16: Note 17:

Free CN by Epstein colorimetric method [45]. Colorimetric method [67]. Tartaric acid distillation, followed by colorimetric method [68]. Pyridine-pyrazalone method [69]. Initial concentration verified based on titration with silver nitrate, using rhodamine indicator. Light spectrophotometry (Way, J.L., personal communication). Lambert et al. [70]. Mertens [71]. Pyridine-barbituric acid method with pH and temperature correction [67,72] . Flame spectrophotometry (after filtering). Weak acid dissociable method 4500-CN method I [73]. APHA [74]. Pyridine-pyrazolone spectrophotometric method [75]. Pyridine-barbituric acid method [74]. Spectrophotometric analysis sensitive to 20 µg/l in sea water [76]. Cyanide electrode. Colorimetric [77].

a Analytical notes: U = unmeasured, M-NR = measured, but not reported, NR = not reported. Detailed numeric notes are as follows:

Protozoan Spirostomidae Spirostomum ambiguum (protozoan)

Pocilloporidae Pocillopora damicornis (coral) Stylophora pistillata (coral)

Moinidae Moinodaphnia macleayi (water flea)

Hydridae Hydra viridissima (hydra)

815

520,272

5203

67 >200

>200 >200

[92]

[88]

[42]

[91] [91]

[91] [91]

Aquatic Toxicity of Cyanide 265

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 265 — #15

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 266 — #16 HCN HCN NR

LC LC LC

HCN HCN

HCN

LC

LC LC

NR

NaCN NaCN NaCN

LC LC LC ELS

NaCN

Cyanide form

ELS

Test type

29–51-d

83-d 83-d

112-d 112-d

144-d

28-d

28-d 56-d 256-d

30-d

Exposure duration

ChV (NOEC-LOEC) ChV (NOEC-LOEC) ChV (NOEC-LOEC)

Spectophotometric analysis

ChV (NOEC-LOEC) ChV (NOEC-LOEC)

ChV (NOEC-LOEC)

ChV (NOEC-LOEC)

ChV (NOEC-LOEC) ChV (NOEC-LOEC) ChV (NOEC-LOEC)

ChV (NOEC-LOEC)

Endpointb

Note 1 Note 1

Note 1 Note 1

Note 1

NR

Note 1 Note 1 Note 1

Note 1

Analytical methoda

Mortality

Growth Reproduction

Growth Reproduction

Reproduction

Mortality

Growth Growth Reproduction

Mortality

Effect

a Analytical notes: U = unmeasured, M-NR = measured, but not reported, NR = not reported. Note 1: Free CN by Epstein colorimetric method [45]. b ChV = chronic value, calculated as the geometric mean of the NOEC and LOEC (values given in parantheses).

Gammaridae Gammarus pseudolimnaeus (amphipod) Mysidae Americamysis bahia (mysid)

Invertebrate Asellidae Asellus communis (isopod)

Cyprinodontida Cyprinodon variegatus (sheepshead minnow) Salmonidae Salvelinus fontinalis (brook trout)

Fish Centrarchidae Lepomis macrochirus (bluegill sunfish) Cyprinidae Pimephales promelas (fathead minnow)

Species

TABLE 14.3 Chronic Toxicity of Cyanide to Aquatic Animals (Acceptable According to USEPA [3])

[62] [62] [97]

69.71 (43–113)

[62] [62]

[96]

[95]

[94] [94] [94]

[93]

Reference

25.92 (21–32) 18.33 (16–21)

34.06 (29–40) 34.06 (29–40)

7.849 (5.6–11.0)

36.12 (29–45)

24 (21.2–27.1) 58.5 (53.9–63.6) 16.39 (13.3–20.2)

13.57 (9.3–19.8)

Concentration (µg CN/l)b

266 Cyanide in Water and Soil

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 267 — #17 U

U

60-h NR 24-h NR

96-h

NaCN KCN KCN KCN

NR

NR U

Population growth

Population growth

EC50

Population growth Mortality

Decreased potassium uptake

LOEC

LOEC LC90

—

EC50

U

32-d

NR

U

Enzyme inhibition

—

—

—

—

Root weight

Respiration inhibition

—

NR

NR

NR

Growth rate

Reduced tetrasporophyte growth Reduced tetrasporophyte growth Reduced female growth Stopped sexual reproduction

Population growth Population growth

Effect

NOEC

— — — —

EC50 LOEC

Endpoint

U

M-NR M-NR M-NR M-NR

M-NR U

Analytical methoda

10-d

14-d 14-d 14-d 14-d

72-h 72-h

Exposure duration

KCN

KCN KCN KCN KCN

NaCN NaCN

Cyanide form

a Analytical notes: U = unmeasured, M-NR = measured, but not reported, NR = not reported.

Chlamydomonadaceae Chlamydomonas sp. (green algae) Chlorellaceae Prototheca zopti (green alga) Chlorophyceae Chlorella sp. (green alga) Haloragaceae Myriophyllum spicatum (Eurasian watermilfoil) Lemnaceae Lemna gibba (duckweed) Microcystaceae Microcystis aeruginosa (blue-green alga) Scenedesmaceae Scenedesmus quadricauda (green alga) Sellaphoraceae Navicula seminulum (diatom)

Champiaceae Champia parvula (red alga)

Algae Bacillariaceae Nitzchia closterium (diatom)

Species

TABLE 14.4 Toxicity of Cyanide to Freshwater Aquatic Plants

277–491

30

75 8000

26,000

22,400

30,000

3000

10–100

16 25 11 11

30 5.3

Concentration (µg CN/l)

[109]

[104, 106–108]

[102–104] [105]

[101]

[100]

[99]

[98]

[16]

[40] [40] [40] [40]

[41] [41]

Reference

Aquatic Toxicity of Cyanide 267

CYWS “L1666_C014” — 2005/10/21 — 17:00 — page 268 — #18

Cancer magister (Dungeness crab) Cancer oregonensis (Oregon cancer crab)

Cancer irroratus (rock crab)

Calyptraeidae Crepidula fornicata (Atlantic slippershell) Cancridae Cancer gracilis (slender crab)

Invertebrate Acartiidae Acartia clausi (copepod) Ampeliscidae Ampelisca abdita (amphipod)

(winter flounder)

Pleuronectidae Pseudopleuronectes americanus

Cyprinodontidae Cyprinodon variegatus (sheepshead minnow)

Fish Atherinidae Menidia menidia (Atlantic silverside)

Species

Zoeae Zoeae Larva Larva Zoeae Zoeae Zoeae Zoeae

NR

NR NR NR NR

1-d old larvae

NR

NR

Age/size

96-h

96-h 96-h 96-h 96-h 96-h 96-h 96-h 96-h

NaCN NaCN KCN KCN NaCN NaCN NaCN NaCN

96-h 96-h 96-h 96-h

96-h

96-h

96-h

Exposure duration

NR

KCN NR NR NR

NaCN

NR

KCN

Cyanide form

Note 11 Note 11 M-NR M-NR Note 11 Note 11 Note 11 Note 11

U

U U U U

U

M-NR

M-NR

Analytical methoda

LC50 LC50 LC50 LC50 LC50 LC50 LC50 LC50

LC50

LC50 LC50 LC50 LC50

LC50

LC50

LC50

Endpoint

TABLE 14.5 Acute Toxicity of Cyanide to Seawater Aquatic Animals (Acceptable According to USEPA [3])

Mortality Mortality Mortality Mortality Mortality Mortality Mortality Mortality

Mortality

Mortality Mortality Mortality Mortality

Mortality

Mortality

Mortality

Effect

153 135 4.2 5.7 51 92 111 154

>10,000

30 1220 1150 704

372

300

59

Concentration (µg CN/l)

[36] [36] [35] [35] [36] [36] [36] [36]

[110]

[112] [113] [113] [113]

[111]

[95]

[110]

Reference

268 Cyanide in Water and Soil

Post-larvae

Penaeidea Penaeus monodon (banana prawn) NaCN

NaCN

NR KCN KCN

NaCN NaCN

NaCN NaCN

96-h

48-h

96-h 96-h 96-h

96-h 96-h

96-h 96-h

Note 16

Note 16

Note 15 U U

U U

Note 11 Note 11

LC50

EC50

LC50 LC50 LC50

LC50 LC50

LC50 LC50

Note 1: Note 2: Note 3: Note 4: Note 5: Note 6: Note 7: Note 8: Note 9: Note 10: Note 11: Note 12: Note 13: Note 14: Note 15: Note 16: Note 17:

Free CN by Epstein colorimetric method [45]. Colorimetric method [67]. Tartaric acid distillation, followed by colorimetric method [68]. Pyridine-pyrazalone method [69]. Initial concentration verified based on titration with silver nitrate, using rhodamine indicator. Light spectrophotometry (Way, J.L., personal communication). Lambert et al. [70]. Mertens [71]. Pyridine-barbituric acid method with pH and temperature correction [67,72]. Flame spectrophotometry (after filtering). Weak acid dissociable method 4500-CN method I [73]. APHA [74]. Pyridine-pyrazolone spectrophotometric method [75]. Pyridine-barbituric acid method [74]. Spectrophotometric analysis sensitive to 20 µg/l in sea water [76]. Cyanide electrode. Colorimetric [77].

Mortality Mortality Mortality

Mortality Mortality

Mortality Mortality

Mortality

Development

a Analytical notes: U = unmeasured, M-NR = measured, but not reported, NR = not reported. Detailed numeric notes are as follows:

Larva

NR NR

Post-larvae

Juvenile Juvenile

Zoeae Zoeae

Pectinidae Chlamys asperrimus (doughboy scallop)

Americamysis bigelowi (mysid)

Mysidae Americamysis bahia (mysid)

Dinophilidae Dinophilus gyrociliatus (archiannelid)

Cancer productus (red crab)

110

28.6

113 93* 124

3152 4021

219 107

[12]

[11]

[97,115] [112] [112]

[114] [114]

[36] [36]

Aquatic Toxicity of Cyanide 269

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Cyanide in Water and Soil

270

conducted since 1984. No new studies on cyanide toxicity to amphibians, algae, or aquatic plants, published after the 1984 AWQC, were identified. All toxicity data are reported as µg CN/l, unless otherwise specified. If data were reported in the original study in units of µg HCN/l, endpoints were converted to µg CN/l, based on the M ratio of CN− to HCN.

14.3.1 ACUTE LETHALITY TO FRESHWATER FISH Acceptable acute cyanide toxicity data were available for 17 species from six taxonomic families of freshwater fish (Table 14.1). The rainbow trout (Oncorhynchus mykiss) was the most acutely sensitive freshwater organism, with 96-h LC50 values ranging from 27 to 97 µg CN/l (geometric mean of 59 µg CN/l). Cyanide appeared to be very fast acting, because significant mortality was observed in rainbow trout exposed to 140 µg CN/l for less than 8 h [13]. In comparision, other salmonids are slightly less sensitive than rainbow trout. For example, the 96-h LC50 for Atlantic salmon (Salmo salar) was 90 µg CN/l, and the 96-h LC50 for brook trout (Salvelinus fontinalis) ranged from 52 to 507 µg CN/l (geometric mean of 105 µg CN/l; Table 14.1). The brook trout toxicity data also demonstrated that swim up fry and juveniles are much more sensitive than sac fry. The geometric mean LC50 values for juveniles and swim up fry were 85 and 84 µg CN/l, respectively, while the LC50 for sac fry was 266 µg CN/l. In general, warmwater fish species are less sensitive than salmonids, although some warmwater species are comparable in sensitivity. For example, within the family Percidae, 96-h LC50 values for the yellow perch (Perca flavescens) ranged from 89 to 330 µgCN/l (geometric mean of 139 µg CN/l; Table 14.1). As with brook trout, a life stage effect was observed, with juveniles being much more sensitive than embryos or fry. The geometric mean LC50 was 94 µg CN/l for juveniles, and 288 and 315 µg CN/l for embryos and fry. Likewise, the most sensitive species and life stage within the family Centrarchidae were juvenile bluegill sunfish (Lepomis macrochirus), with a geometric mean LC50 of 98 µg CN/l. Thus, based on the data for juveniles, yellow perch and bluegill sunfish are comparable in sensitivity to swim up fry and juvenile brook trout. Other warmwater fish species are less acutely sensitive than salmonids, yellow perch, or bluegills. For example, within the family Cyprinidae, the fathead minnow (Pimephales promelas) had the lowest LC50 values, with a geometric mean LC50 of 131 µg CN/l. Within the family Poecilia, the guppy (Poecilia reticulate) was comparable in sensitivity to the fathead minnow, with a single LC50 value of 147 µg CN/l (Table 14.1). When multiple acute tests were available for a given species, 96-h LC50 values did not vary widely, with nearly all values falling within two standard deviations of the species mean (the variation is even less if life stage is accounted for). The least amount of variation was found for rainbow trout, for which all the 96-h LC50 values fell within one standard deviation of the species mean. A similar pattern was found for each family. The minimal variation in LC50 values within each family suggests that the potential toxicity of cyanide to species within the families Salmonidae, Centrarchidae, Percidae, and Cyprinidea — may be estimated with confidence. In fact, the available data suggest that the majority of fish species may have a similar sensitivity to cyanide, with geometric mean LC50 values for most species ranging from 59 to 330 µg/l (when accounting for the most sensitive life stage). However, a few species of freshwater fish appear to be uniquely tolerant of cyanide exposure. For example, tropical warmwater cyprinids and ciclids may be even more resistant to cyanide, with LC50 values ranging from 175 to 1051 µg CN/l (Table 14.1). Although no standard toxicity test data are available, the brown bullhead (Ictalurus nebulosus), a particularly hardy species in the Ictaluridae family [14], tolerated 1600 µg CN/l for up to 8 h. Species within this family are well known for having a higher tolerance to toxic substances and anoxic conditions than many freshwater fish [14]. It is thus possible that other species of fish in the family Ictaluridae may also demonstrate a similar tolerance to cyanide, but data enabling a test of this hypothesis are not currently available.

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14.3.2 ACUTE TOXICITY TO FRESHWATER INVERTEBRATES Studies evaluating the acute toxicity of cyanide towards freshwater invertebrates have been conducted with 21 species comprising 17 taxonomic families (Table 14.1). While cyanide was not as acutely toxic to freshwater invertebrates as it is to many of the fish species, two studies conducted using the freshwater crustacean Daphnia pulex reported 48- and 96-h LC50 values of 110 and 83 µg CN/l, respectively [15,16]. A study with Daphnia magna resulted in a 96-h LC50 value of 90 µg CN/l [17]. These LC50 values are comparable to those reported for several sensitive fish species (Section 14.3.1). Other 96-h LC50 values for unidentified Daphnia spp. were higher, ranging from 133 to 360 µg CN/l. While only limited acute toxicity data were available for aquatic insects, these data showed that they were less sensitive to cyanide than Daphnia spp., with 96-h LC50 values ranging from 228 CN/l for the water scorpion (Ranatra sp.), to 2,490 µg CN/l for the midge Tanytarsus dissimilis. While the LC50 values for daphnids and insects did not show high variation, there were large differences in the 96-h LC50 values for the two amphipod species tested within the family Gammaridae. Gammarus pseudolimnaeus was about 5 times more sensitive to cyanide than G. fasciatus, with 96-h LC50 values of 167 and 900 µg CN/l, respectively. Similar results were observed for the two species of gastropods tested in the families Planorbidae and Physidae. The snail Helisoma trivolvis demonstrated extreme tolerance to cyanide, with a 96-h LC50 value >50,000 µg CN/l, while the snail Physa heterostropha demonstrated a 96-h LC50 value of 432 µg CN/l which is similar to those observed in aquatic insects. Such large ranges in toxicity values are unusual for closely related species, and so these results may warrant closer examination.

14.3.3 ACUTE SUBLETHAL EFFECTS ON FRESHWATER FISH Acute sublethal endpoints are defined here as nonlethal effects observed in test organisms exposed to cyanide for short-term durations (≤96-h). Owing to cyanide’s mode of action (Section 14.1), its effects on an organism often rapidly appear when present in sufficient concentrations. When cyanide concentrations are below lethal concentrations, however, sublethal effects (e.g., biochemical, physiological) may also occur over time periods shorter than those typically considered “chronic” (i.e., one week or greater). A number of acute sublethal tests have been conducted with rainbow trout, brown bullhead (I. nebulosis) and elephant nose fish (Gnathonemus petersi; Table 14.2). The most sensitive acute sublethal endpoint was observed in rainbow trout, for which a 15-min exposure of gametes to cyanide at 0.4 µg CN/l was found to significantly impair fertility. Other endpoints tested include tissue damage, and changes in hormone levels, blood chemistry, respiration, and equilibrium (Table 14.2). For the most part, once removed from the presence of cyanide, an organism receiving less than an acute dose should recover, because cyanide is detoxified to SCN− . In some cases, however, some sublethal effects can exert lasting (i.e., one week or greater) negative impacts on the exposed organism [18]. For example, significant gill damage consisting of severe necrotic damage and desquamation was noted in G. petersi after exposure to 50 µg CN/l for 6 h. Exposure to 100 µg CN/l cyanide resulted in even greater damage, with gill tissues being reduced to the supporting skeleton in many places. These results suggest that adverse effects may persist over time, owing to acute gill injury. Sublethal cyanide concentrations may also impair fish reproduction (see also Section 14.3.4). For example, blood levels of the hormone cortisol are considered indicators of stress in both Cyprinidae [19] and Salmonidae [21], as are changes in blood cholesterol [22,23]. Both of these compounds were significantly different in rainbow trout exposed to 52 µg CN/l cyanide for 1 to 3 days. Induction of these stress responses could alter various steps in fish reproduction [24]. Other sublethal endpoints that respond to even short-term cyanide exposures include respiration and swimming activity. Respiration of rainbow trout and brown bullheads was significantly reduced in adults exposed to cyanide at 70 and 700 µg CN/l, respectively, for 3 to 4 h. Significant reductions in

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swimming activity, or rather cessation of swimming activity, was observed in rainbow trout exposed to 60 µg CN/l for 3 h. Significant reductions in the swimming activity of brown bullhead were observed at 200 µg CN/l; however, an exposure period of only 30 min was required to induce the observed effect.

14.3.4 CHRONIC TOXICITY TO FRESHWATER FISH Chronic cyanide data considered acceptable for use in AWQC derivation were available for four species of freshwater fish in the families Salmonidae, Centrarchidae, and Cyprinidae (Table 14.3). As with the acceptable acute toxicity data (Section 14.3.1), the variability in chronic sensitivities between species is not very wide. For example, the chronic values (geometric mean of NOEC [no observable effects concentration] and LOEC) for brook trout (S. fontinalis), fathead minnow (P. promelas), and bluegill sunfish (L. macrochirus) were 7.8, 16.4, and 13.6 µg CN/l, respectively. For all three species, the most sensitive chronic endpoint was reproduction. No toxicity data from rainbow trout early life stage tests were available, but subchronic data for a variety of endpoints suggest a chronic sensitivity similar to other fish. The sensitivity of fish reproduction to chronic cyanide exposures may be deduced from studies that evaluated key reproductive hormones and structures in Salmonidae exposed to low concentrations of cyanide (Table 14.2). For example, vitellogenin is a glycolipophosphoprotein produced in nonmammalian female vertebrates during the process of yolk formation [25], and its production can be inhibited by cyanide. Studies by Ruby et al. [25] and Szabo et al. [26] reported significant reductions in the amount of vitellogenin in the plasma of adult female O. mykiss exposed over a 12-day period to 10 µg CN/l. Only 5 µg CN/l was required to do the same in Salmo salar ouananiche [27]. Blood levels of the hormone dopamine, which has been identified as an inhibitory factor in the release of gonadotropin [28,29] was also found to be significantly elevated in rainbow trout exposed to cyanide [26]. Additionally, the size of the female trout’s gonads in relation to body size (gonadosomatic index) was also reduced by 10 µg CN/l over the same time period, when exposed to 10 µg CN/l [25]. A similar reproductive impairment was found in male O. mykiss exposed to 10 µg CN/l over 12 days. These male fish had a significantly different number of spermatogonial cysts, compared to the control group [26,30]. The findings of another series of studies conducted at Oregon State University between 1966 and 1973 with salmonid fishes help illustrate the effects of cyanide on chronic endpoints not typically studied, including locomotor and feeding behavior. Leduc [31] exposed juvenile coho salmon (Oncorhynchus kisutch) to cyanide for up to 24 days, and assessed the effects on survival, growth, and feeding behavior. The results of this study are not conclusive because of disease and excessive control mortalities near test initiation; however, the preliminary results suggested that growth was reduced at 80 µg HCN/l. In the second half of the experiment, salmon exposed to 20 to 80 µg HCN/l grew faster than the controls, suggesting that the fish had adapted to cyanide exposure. Broderius [32] exposed coho salmon yearlings to cyanide for up to 194 h and evaluated locomotor behavior and growth. Fish exposed to concentrations of 10, 30, and 50 µg CN/l were forced to swim against a maximum current of 1.62 ft/sec. The length of time each fish was able to swim against this current was recorded. After the swimming test, fish length and weight were measured. Effects on swimming performance were evaluated after cyanide exposures of 2, 26, 53, 126, and 194 h. After cyanide exposures and swim tests, fish were maintained in cyanide-free solutions for 6, 83, 169, 251, and 337 h, and then their swimming performance was retested. Cyanide exposed fish had a markedly reduced swimming time, compared to control fish. Control fish were able to swim against the maximum 1.62 ft/sec current for a mean of 8.72 min, compared to 3.80, 1.85, and 1.40 min for fish exposed to 10, 30, and 50 µg CN/l, respectively, for 2 h prior to the swimming test. Swimming time improved after recovery in cyanide-free water, but was still reduced, relative to controls. Finally, Negilski [33] exposed juvenile Chinook salmon (O. tshawytscha) in experimental streams to cyanide for two months, and evaluated the effects on biomass, growth, and oxygen consumption.

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A cyanide concentration of 10 µg CN/l in two separate experiments resulted in salmon biomass greater than that measured in the controls. In a third experiment, salmon exposed to 20 µg CN/l exhibited reduced biomass, relative to the control (biomass was approximately 6 g/m2 in the control, compared to approximately 4.3 g/m2 in the salmon exposed to 20 µg CN/l). Overall, these studies suggest that as little as 10 µg CN/l reduces the ability of yearling coho salmon to swim against a current, and a cyanide concentration of 20 µg CN/l may be sufficient to reduce Chinook salmon biomass. These effect concentrations are consistent with other chronic and sublethal test results for salmonids (Tables 14.2 and 14.3), and so salmonid swimming behavior is not uniquely sensitive to cyanide exposure, when compared to more typically studied chronic endpoints (e.g., survival, growth, and reproduction).

14.3.5 CHRONIC TOXICITY TO FRESHWATER INVERTEBRATES Chronic toxicity studies with freshwater invertebrates considered acceptable for AWQC derivation have only been conducted with one isopod (Asellus communis) and one amphipod (Gammarus pseudolimnaeus; Table 14.3). Reproduction tended to be the most sensitive chronic endpoint, with G. pseudolimnaeus being the more sensitive of the two invertebrates tested with a reproductive LOEC of 21 µg CN/l. Since publication of USEPA’s 1984 AWQC, two additional freshwater invertebrate chronic studies have been conducted with the cladoceran Moinodaphnia macleayi and the hydra Hydra viridissima (Table 14.2). Even the most sensitive reproductive LOEC for M. macleayi (67 µg CN/l) was less than those for the two tests conducted using the amphipod and the isopod.

14.3.6 FRESHWATER ALGAE AND VASCULAR PLANTS The mechanism of cyanide toxicity in plants is similar to that of animals, and is mediated through complexation with iron in the enzyme cytochrome oxidase [34]. As with animals, this inhibits cellular respiration, resulting in reduced ATP production. Without sufficient ATP levels, plants are incapable of conducting ion uptake, and phloem transport can be inhibited. If these functions are sufficiently impaired, the plant will eventually die [34]. In the 1984 AWQC [2], cyanide toxicity data were identified for seven species of freshwater algae and vascular plants; no newer studies were identified (Table 14.4). The most sensitive endpoint tested was growth inhibition, which occurred in the cyanobacterium Microcystis aeruginosa and the green algae Scenedesmus quadricauda, exposed to 75 and 30 µg CN/l, respectively. In the diatom Navicula seminuium, a 50% reduction in division rates was observed from 277 to 491 µg CN/l. In contrast, aquatic vascular plants were very resistant to cyanide. Potassium uptake was decreased in duckweed (Lemna gibba) exposed to 26,000 µg CN/l, and Eurasian water milfoil (Myriophyllum spicatum) exposed to 22,400 µg CN/l had significantly lower root weights than control plants. These studies suggest that cyanide is more toxic to algae than vascular plants. The sensitivity of algae to cyanide is also within the range of that observed in many freshwater invertebrates, but less than that of the most acutely sensitive freshwater fish.

14.4 TOXICITY TO MARINE BIOTA Cyanide toxicity to marine species has been studied less extensively than in the case of freshwater species. As we did for freshwater biota (Section 14.3), our primary source of aqueous toxicity data was the USEPA’s AWQC for cyanide [2], which was supplemented with more recent studies from the primary scientific literature. All toxicity data are reported here as µg CN/l, unless otherwise specified.

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14.4.1 ACUTE TOXICITY TO MARINE FISH Acceptable acute toxicity studies have only been conducted with three species of marine fish, with Menidia menidia (Atlantic silverside) being the most acutely sensitive, with a 96-h LC50 value of 59.3 µg CN/l (Table 14.5). This LC50 value is within the range of the most acutely sensitive freshwater species, O. mykiss (Table 14.1). Pseudopleuronectes americanus (winter flounder) and Cyprinodon variegatus (sheepshead minnow) were less sensitive to cyanide, with 96-h LC50 values of 372 and 300 µg CN/l, respectively, which are similar to several of the less sensitive freshwater species (e.g., Salvelinus fontinalis, Perca flavescens, and Pimephales promelas). These studies demonstrate that marine and freshwater fish species have similar sensitivity to cyanide, which is not surprising, given the common mode of toxic action for cyanide to most organisms (Section 14.1).

14.4.2 ACUTE TOXICITY TO MARINE INVERTEBRATES Acceptable acute toxicity data were available for 16 marine invertebrate species in 11 families (Table 14.5). Perhaps the most important studies on acute cyanide toxicity to marine invertebrates are those using crabs within the genus Cancer. The most acutely sensitive species to cyanide was Cancer irroratus (Eastern rock crab), with 96-h LC50 values ranging from 4.2 to 5.7 µg CN/l [35]. But studies with West Coast Cancer species found them to be over 10 times more tolerant of cyanide than C. irroratus [36]. The commercially important C. magister (Dungeness crab) had the lowest 96-h LC50 value of the west coast species (51 and 92 µg CN/l), while the remaining three West Coast species tested had 96-h LC50 values ranging from 107 to 219 µg CN/l (Table 14.4). Therefore, it is possible that the genus Cancer is not as acutely sensitive to cyanide as the original C. irroratus study suggested. To further assess the differences in sensitivity between East and West Coast Cancer spp., the C. irroratus study is currently being repeated, and preliminary results suggest that its acute sensitivity to cyanide may instead be more similar to that of C. magister [37]. The 96-h LC50 values for most other crustaceans were similar to those for the West Coast Cancer species (Table 14.5). For example, LC50 values between 93 and 124 µg CN/l were observed for Mysidopsis bahia and M. bigelowi (both since reclassified under the genus Americamysis), respectively, and a 96-h LC50 value of 110 µg CN/l was observed with Penaeus monodon (tiger prawns). The copepod Acartia clausi was more sensitive to cyanide than the West Coast Cancer species, with a 96-h LC50 value of 30 µg CN/l. The doughboy scallop exhibited acute sensitivity similar to that of Acartia, with a 96-h EC50 value for normal embryo development of 28.6 µg CN/l. The sensitivity of C. asperrimus larvae to cyanide is not surprising, given that bivalve developmental tests are typically quite sensitive to chemicals [38]. The benthic amphipod Ampelisca abdita, however, was much more tolerant to cyanide than the other organisms tested from the subphylum Crustacea, with an LC50 value of 704 µg CN/l. Studies on the sublethal toxicity of cyanide to saltwater invertebrates following short-term exposures were limited to several coral species (Table 14.2). Overall, corals do not appear to be very sensitive to cyanide, with NOECs ranging from 260 µg CN/l, based on change in fluorescence, to 520,272 µg CN/l, based on inhibition of photosynthesis in symbiotic algae.

14.4.3 CHRONIC TOXICITY TO MARINE FISH AND INVERTEBRATES Chronic toxicity data acceptable for use in AWQC derivation were only identified for a single marine fish, the sheepshead minnow C. variegatus, and a single marine invertebrate, the mysid shrimp A. bahia (Table 14.3). The sheepshead minnow was more chronically sensitive to cyanide, with a mortality LOEC of 45 µg CN/l, while the mysid shrimp’s mortality LOEC was over 2-fold higher, at 113 µg CN/l. In contrast, a chronic test conducted using the bivalve Scapharca inaequivalvis (acrid blood clam), found them to be very tolerant to long-term exposure (11.2-d) to high concentrations (26,000 µg/l) of cyanide. This species of clam comes from a primitive family of clams (Arcidae)

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that have blood cells containing hemoglobin instead of the hemocyanin or myoglobin found in most bivalves [39]. This allows these species to extract oxygen from the water with greater efficiency than bivalves using other respiratory pigment [39], and thus, this may help explain why S. inaequivalvis is so tolerant of high concentrations of cyanide.

14.4.4 MARINE ALGAE AND VASCULAR PLANTS Cyanide toxicity data are available for several species of marine algae, but no data are available for marine vascular plants (Table 14.4). The most sensitive endpoints identified were associated with reproduction in the red algae Champia parvula. Growth of female C. parvula was significantly reduced when exposed to 11 µg CN/l, as was sexual reproduction [40]. Two species of marine green algae were substantially less sensitive, with respiration in Prototheca zopil being significantly inhibited at 3,000 µg CN/l, and enzyme inhibition in Chlorella sp. occurring at 30,000 µg CN/l (Table 14.4). In algal tests conducted since the USEPA AWQC, growth of the diatom Nitzschia closterium was significantly reduced when exposed to 10µg CN/l for 72 h, with a 50% reduction in growth at 57 µg CN/l [41]. Algae associated with corals appeared to be much less sensitive to cyanide, with at least 2602 µg CN/l needed to significantly reduce the number of zooxanthellae in the coral species Plesiastrea versipora and Pocillopora damicornis [42,43]. Therefore, at least for the red alga C. parvula and the diatom N. closterium, marine algae can be even more sensitive to cyanide than the most acutely and chronically sensitive marine animal species. However, as reviewed in Section 14.5, current AWQC for cyanide would still be protective of growth and reproduction in even these species.

14.5 REGULATORY CRITERIA FOR PROTECTION OF AQUATIC LIFE In this section, we review current regulatory thresholds for protection of aquatic life from cyanide exposure in the United States, and whether the new data reviewed above might warrant changes in these threshold concentrations. Several methods are available for deriving water quality criteria that can be used to establish the maximum levels of contaminant exposure below which there should not be significant harm to aquatic biota. Some of the most widely accepted and detailed methods are those used by the USEPA to derive AWQC for the protection of aquatic life [3]. These methods were developed for setting water quality criteria and standards for compliance with the Clean Water Act (CWA), but they are also useful in situations where strict CWA compliance is not the immediate concern, or where no officially promulgated criterion is available. National AWQC set maximum threshold concentrations of contaminants for both freshwater and marine environments. These criteria are derived from empirical toxicity data, and are designed to be stringent enough to protect most sensitive species potentially exposed to a contaminant in any water body in the United States. Below these thresholds, no adverse effects are anticipated. The thresholds derived in each AWQC are designed to protect all but the most sensitive 5% of species. If data suggest that any commercially or recreationally important species are not protected at this level, then these values can be adjusted to provide sufficient protection for these species as well. AWQC are derived following the guidelines and terminology used by USEPA [3]. The first step is to compile acute and chronic toxicity data that meet certain guidelines to ensure scientific acceptability. For each species with acceptable acute toxicity data, the species mean acute value (SMAV) is calculated as the geometric mean of available 48 to 96-h LC/EC50 values for that species. The genus mean acute value (GMAV) is then calculated as the geometric mean of available SMAVs for each genus. The 5th percentile of the GMAVs is identified as the final acute value (FAV), which is divided by two to determine the criterion maximum concentration (CMC), more commonly termed

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TABLE 14.6 Summary of Four Most Sensitive GMAVs Used to Derive Current AWQC Freshwater Genus

Saltwater

GMAV (µg CN/l)

Genus

GMAV (µg CN/l)

Lepomis Perca Salvelinus Salmoa

99.28 92.64 85.80 63.45

Mysidopsisb Menidia Acartia Cancer

118.4 59 30 4.893

FAV CMC ACR FCV/CCC

44.73c 22.36 8.568 5.221

FAV CMC ACR FCV/CCC

2.030 1.015 NA 1.015

Remaining acronyms are defined in text. a Salmo includes rainbow trout, which has since been reclassified under Oncorhynchus. b Mysidopsis has since been reclassified as Americamysis. c The FAV based on the 5th percentile of GMAVs was 62.68, but this was lowered to 44.73

to be protective of rainbow trout (a recreationally and economically important species). NA = Not applicable (see text).

the “acute criterion.” It is important to note that the 5th percentile is calculated based solely on the four most sensitive GMAVs and the total number of GMAVs [3]. The chronic criterion may be derived in a manner similar to the CMC, but chronic toxicity data for a sufficient number of species are typically unavailable (as is the case for cyanide). It is typically necessary to apply an acute–chronic ratio (ACR) to the FAV to estimate the final chronic value (FCV). Unless other data are available to suggest the FCV is under-protective of the aquatic community, the criterion continuous concentration (CCC), or chronic criterion, is set equal to the FCV.

14.5.1 BASIS OF CURRENT AWQC The current USEPA acute and chronic AWQC for freshwater systems are 22.36 and 5.221 µg CN/l, respectively [2]. They calculated 15 GMAVs, the four lowest of which are presented in Table 14.6. As shown, the most sensitive genera are fish species, including two genera within family Salmonidae. Because the FAV calculated as the 5th percentile of the GMAVs was greater than the SMAV for rainbow trout, the FAV was reduced to the rainbow trout SMAV because it is a recreationally and economically important species. Applying the ACR of 8.568 to this lower FAV resulted in a FCV of 5.221. For saltwater species, the current acute and chronic AWQC are both 1.015 µg CN/l which is substantially lower than the freshwater criteria concentrations [2]. The USEPA calculated eight GMAVs, of which the four lowest are shown in Table 14.6. Unlike the freshwater toxicity data, three of the four most sensitive GMAVs are for invertebrates, with the East Coast rock crab (Cancer irroratus) GMAV being approximately six times lower than the next lowest GMAV. Driven largely by the low Cancer GMAV and the small number of GMAVs available, the FAV was calculated to be 2.030 µg CN/l, resulting in a CMC of 1.015. Because ACRs were only available for insensitive species, the FCV was set equal to the CMC under the premise that the C. irroratus larval development study probably provided a better indication of its chronic sensitivity than usage of an ACR.

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TABLE 14.7 Summary of Four Most Sensitive GMAVs and Re-Calculation of AWQC Using an Updated Toxicity Database Freshwater

Saltwater

GMAVa (µg CN/l)

Genus

GMAVa (µg CN/l)

Perca Salmo Salvelinus Oncorhynchus

92.64 90 85.80 46.27

Americamysis Cancer a Menidia Acartia

118.4 62.64 59 30

FAV CMC ACR FCV/CCC

46.27b 23.14 8.568 5.400

FAV CMC ACR FCV/CCC

20.42 10.21 6.45c 3.17

Genus

Remaining acronyms are defined in text. a The GMAV includes data from C. irroratus [35] and all four West Coast Cancer species [36]. b The FAV based on the 5th percentile of GMAVs was 55.64, but this was lowered to 46.27 to be protective of rainbow trout (a recreationally and economically important species). c ACR and its basis from Brix et al. [36].

14.5.2 COMPARISON OF CURRENT AWQC TO UPDATED TOXICITY LITERATURE Using the updated toxicity data presented in Sections 14.3 and 14.4, water quality criteria can be recalculated from the studies considered scientifically acceptable for use in criteria derivation [3]. For freshwater species, these new studies would exert little influence on the magnitude of the CMC. Toxicity data were identified for eight new genera, and one species (rainbow trout) was taxonomically reclassified, effectively adding another genera, resulting in toxicity data for 24 genera (compared to 15 in the 1984 AWQC). However, no species more sensitive than the rainbow trout were identified, so a revised FAV (and CMC) would still be based on the SMAV for rainbow trout (Table 14.7). A revised CMC would be slightly different than the 1984 CMC, because an additional rainbow trout study modified the SMAV. Since no new studies were published from which an ACR could be derived, the CCC would also change only slightly (Table 14.7). Although only one additional saltwater genus which met the USEPA guidelines has been tested since 1984, revised saltwater criteria could show substantial changes. The most important of these new studies is a reevaluation of cyanide toxicity to four West Coast species of rock crab [Cancer spp., 36]. A GMAV based just on these four West Coast species would be 118.5 µg CN/l, which is about 24 times greater than the 4.893 µg CN/l GMAV for C. irroratus currently reported in the 1984 AWQC (see also Table 14.5). Preliminary results from new acute studies with C. irroratus suggest its acute sensitivity to cyanide may instead be more similar to C. magister [37], so a higher GMAV for Cancer may indeed be a more accurate reflection of its sensitivity to cyanide. If all of the values from the currently published Cancer data are averaged together, the ranking of the four most sensitive genera would change significantly, and the FAV and CMC would increase approximately 10 times to 20.42 and 10.21 µg CN/l, respectively (Table 14.7). Because Cancer spp. would no longer drive the CMC, a FCV could be calculated using an ACR. Using the ACR of 6.45 recommended in Brix et al. [36], the FCV would be 3.17 µg CN/l.

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14.5.3 CRITERIA IMPLEMENTATION AS A FUNCTION OF CYANIDE CHEMICAL FORM Given that free, rather than complex, cyanides are most relevant to understanding the bioavailability and toxicity of cyanide, it is important to carefully consider the chemical form of cyanide to be used for compliance monitoring in effluents and receiving waters. In the 1984 AWQC, criteria concentrations were derived on the basis of free cyanide, but it was recommended that the criteria should instead be implemented on the basis of total cyanide [2]. This is because no acceptable standardized methods were available for measurement of free cyanides at the time. Upon closer examination of the studies used in the 1984 AWQC, its derivation on the basis of free cyanide is based on two assumptions. First, because most studies were conducted by spiking simple cyanide (typically KCN or NaCN) into simple inorganic test solutions, it was generally assumed that simple cyanides would have dominated cyanide speciation. While this may be a reasonable assumption in freshwaters, it is possible that small but significant amounts of metal cyanides can form in some natural seawaters used in toxicity testing [44]. Second, even though most of the studies used in cyanide criteria derivation used analytical measurements of free cyanide, it is not clear to what extent the methods used would have reliably measured free cyanides, when compared to more current standard methods. In the majority of the studies used in the derivation of the 1984 AWQC for cyanide [2], free cyanide was determined using the pyridine–pyrazalone colorimetric method from Standard Methods [45], with direct colorimetric development and no distillation. This was true at least for all of the four most acutely sensitive freshwater tests, and three of the four most acutely sensitive seawater tests used in FAV calculations, and for all of the chronic studies used in FCV calculations [2]. It is important to note that these colorimetric methods may exhibit more analytical uncertainties or interferences than more recently developed methods that measure weak-acid dissociable (WAD) or free cyanide by micro diffusion (see Chapter 7). However, empirical comparisons to more current standardized methods suggest they are still likely to be reasonably accurate predictors of toxicity as a function of free cyanide, at least in freshwaters [44]. A significant number of the remaining studies reviewed in the 1984 AWQC either did not measure free cyanide, or did not measure any form of cyanide in test solutions. Given the importance of free cyanide to quantifying its toxicity, and because of the volatility of HCN under some conditions, this would appear to challenge the accuracy of any regulatory criteria based on tests that did not adequately verify free cyanide concentrations. However, the studies that did not measure cyanide used insensitive species of little direct numeric importance to criteria derivation, so the lack of analytical verification would have little impact on the accuracy of the AWQC. Studies conducted since the 1984 AWQC have largely used analytical measurements of free cyanide concentrations, so any updates to these criteria would further support their calculation on the basis of free, rather than total, cyanide. For example, newer studies have either used colorimetric measurements without distillation [46], or the more prevalent standard method of WAD cyanide [36], to verify test endpoints on the basis of free cyanide. Coupled with the more widespread use of reliable and standardized free cyanide analytical methods, direct AWQC implementation on the basis of free cyanide concentrations may now be possible.

14.6 SUMMARY AND CONCLUSIONS • Hydrogen cyanide, HCN, acts as a powerful inhibitor of the haemoprotein cytochrome oxidase by preferential binding to the iron porphyrins present in cytochrome oxidase. This binding can cause death by histotoxic apoxia. • While cyanide may exist in a variety of metallocyanide or organic complexes in the aquatic environment, the toxicity of these complexes is largely a function of their dissociation to

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•

•

• •

•

279

free cyanide (i.e., HCN or CN−). As a result, toxicity is best expressed as a function of free, rather than total, cyanide. Cyanide toxicity to aquatic organisms has been reported to vary as a function of water pH, temperature, and dissolved oxygen concentrations, but these relationships are generally not considered to be strong enough to modify toxicity threshold concentrations as a mathematical function of these factors. Rainbow trout (Oncorhynchus mykiss) are the most acutely sensitive freshwater species to cyanide, with a species mean acute toxicity value of 46.3 µg CN/l. Sublethal toxicity can also be expressed as impaired reproduction, growth, or swimming behavior in several other freshwater species, but rarely at concentrations less than 10 µg CN/l. Fewer toxicity data are available for marine animals, and their acute and chronic sensitivity to cyanide is generally similar to that of freshwater species. The only exceptions are marine copepods (Acartia clausi) and Eastern rock crabs (Cancer irroratus), which may be more acutely sensitive than the most sensitive freshwater salmonids. As with freshwater species, endpoints associated with reproduction tend to be the most chronically sensitive. Aquatic plants are sensitive to cyanide at concentrations similar to, or below that of, most aquatic animals, with the most sensitive being the marine alga Champia parvula, which expresses inhibited reproduction at 11 µg CN/l. Existing AWQC for protection of aquatic organisms and their uses in the United States suggest that freshwater organisms would be protected from acute and chronic toxicity at concentrations of 22.4 and 5.2 µg CN/l, respectively. If updated, new toxicity studies would not change these freshwater criteria concentrations substantially. In contrast, new marine toxicity data suggest that the existing AWQC for protection of marine organisms (1.015 µg CN/l) could be 3 to 10 times too conservative for chronic and acute saltwater criteria, respectively. Updated AWQC would also be protective of aquatic plants, and of aquatic dependent wildlife (Chapter 15). Given the advancements in the reliable measurement of free cyanide concentrations, it should be possible to implement updated AWQC on the basis of free, rather than total, cyanide.

ACKNOWLEDGMENTS We thank Diane Lightwood for her assistance obtaining references. We also thank the Water Environment Research Foundation (WERF) project coordinator, Margaret Stewart, and the project subcommittee (Walter Berry, Phillip Dorn Joseph Gorsuch, Jim Pletl, and Mary Reiley) for their review of the manuscript. This work was supported in part by WERF, Project 01-ECO-1.

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of Cyanide to 15 Toxicity Aquatic-Dependent Wildlife Jeremy M. Clark, Rick D. Cardwell, and Robert W. Gensemer CONTENTS 15.1 15.2 15.3 15.4

Distribution of Cyanide in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Exposure Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mechanisms of Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Literature Review Methods and Scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.1 Data Quality Determination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.2 Data Normalization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.2.1 Normalization to mg/kg Body Weight . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.4.2.2 Normalization of Toxic Dose to Cyanide Ion. . . . . . . . . . . . . . . . . . . . . . . . 15.4.2.3 Endpoint Normalization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.5 Data Analysis. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6 Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.1 Route of Exposure — Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.2 Route of Exposure — Birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.3 Comparisons between Birds and Mammals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.3.1 Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.3.2 Food . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.3.3 Direct Injection. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6.4 Simple Cyanide versus Complex Cyanide Compounds . . . . . . . . . . . . . . . . . . . . . . . . . 15.7 Bioaccumulation of Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.8 Toxicity Thresholds for Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.9 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

286 287 287 287 288 288 288 289 289 289 290 290 303 304 304 304 304 304 305 305 306 306 307

The U.S. Environmental Protection Agency’s (USEPA) ambient water quality criteria (AWQC) for cyanide were developed in 1984 [1] and have been used extensively to develop local water quality standards for protection of aquatic life. New knowledge on the relative toxicity of bioavailable cyanide species, and the measurement of cyanide species [2] have prompted a reevaluation of the aquatic toxicity data that serve as the basis of the current national criteria [3; see also Chapter 14]. However, AWQC for protection of aquatic life do not necessarily represent concentrations that would be protective of the entire aquatic ecosystem. Consideration also should be given to the sensitivity of wildlife species whose primary habitats are aquatic or are dependent on aquatic life as a food source. Aquatic-dependent wildlife is comprised of waterfowl, shorebirds (e.g., sandpipers), and aquatic mammals (e.g., otter, beaver). 285

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Here, we review the toxicity of cyanide compounds to aquatic-dependent wildlife exposed via drinking water and food. Our focus was to evaluate the bioavailability1 of cyanide from different exposure pathways and the degree to which toxicity changes when different cyanide compounds pass from the intestinal tract into the bloodstream. More specifically, the purpose of this review is to evaluate the following questions: • What is the relative toxicity of cyanide compounds to aquatic-dependent wildlife? • Does cyanide toxicity to birds and mammals differ materially by route of exposure (e.g., drinking water versus dietary exposure)? • What is the range of toxicity of cyanide compounds and do simple cyanide compounds differ significantly from complex cyanides in toxicity? • Does normalization of the toxic dose of a cyanide compound to free cyanide (HCN and CN− ) concentration provide a more accurate and comparable estimate of a toxicity threshold or reference value? • Which no-effect concentrations appear protective of birds and mammals generally, and aquatic-dependent wildlife specifically? Because toxicologic data for aquatic-dependent wildlife species are extremely limited, data for birds and mammals commonly tested in the laboratory also were used. Testing of surrogate animal species is standard practice in wildlife risk assessments, as relatively few species have been tested, compared to the large number of bird and mammal species of concern [5].

15.1 DISTRIBUTION OF CYANIDE IN THE ENVIRONMENT Cyanide compounds are used for a wide variety of private and industrial processes and formed as a result of certain chemical reactions (Chapter 4). In addition, they are formed naturally by certain plants (Chapter 3); for example, cyanogenic glycosides are produced in cassava and thiocyanate (SCN− ) is produced in plants from the family Brassicaceae [2,6]2 . Anthropogenic sources include mining operations, manufacture of synthetic fabrics and plastics, pesticides, and production intermediates in agricultural chemical production [6,7]. Formation of cyanide compounds during treatment of municipal wastewater can also occur [2; and Chapter 25]. Chemical forms of cyanide in the environment include free cyanide, simple cyanides, metallocyanide complexes, thiocyanate, synthetic nitriles, and organic cyanides (Chapter 2). Free cyanide (CN− and HCN) appears to be the primary toxic form in the aquatic environment [8,9]. In aqueous solution below pH 9.2, the majority of free cyanide exists as hydrogen cyanide, HCN [6]. Simple cyanides typically refer to water-soluble salts of free cyanide such as sodium or potassium cyanide (NaCN and KCN), respectively. In water, NaCN and KCN completely dissociate to produce free cyanide, which is a pH-dependent combination of CN− and HCN [9; and Chapter 5]. Metallocyanide salts produce variable fractions of free cyanide upon dissolution in water, the concentrations of which depend on pH and the metal’s affinity for the CN− ion (e.g., CdCN− , Cu(CN)− 2, − 4− Ni(CN)2− , Zn(CN) , Fe(CN) , etc.; see Chapters 2 and 5). Of the metal–cyanide complexes, iron– 3 4 6 cyanide complexes often predominate in surface waters because of the abundance of iron and the high affinity of CN− for Fe2+ and Fe3+ (Chapter 5). Most environmentally important complexes associated with mining and mineral extraction (e.g., gold) are classified as “weak acid dissociable” (WAD) cyanides [10]. Exceptions are cobalt and iron cyanides, which are not quantified by the WAD cyanide analytical method [2; and Chapters 5 and 7]. 1 The

term “bioavailability” is defined in this context as the degree to which a chemical can be taken up by an organism, subsequently interacting with a biologically important site of action [4]. 2 This family includes cauliflower, cabbage, and turnips.

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Biogenic sources of cyanide consist of various species of bacteria, algae, fungi, and higher plants producing and excreting cyanide compounds (Chapter 3). Elevated concentrations of cyanide occur in many food plants and forage crops, and may represent the greatest sources of cyanide exposure to terrestrial mammals [10]. In this regard, cassava (Manhot esculenta) has received the most study because of its elevated content of organic cyanide compounds (glycosides) and because of its importance as a major food staple in Asia, Africa, South America, and the Caribbean Islands [11].

15.2 EXPOSURE PATHWAYS Animals may be exposed to cyanide or cyanide compounds via a number of pathways. They may ingest food or water containing natural or anthropogenic cyanide. Toxicity from cyanide-producing (cyanogenic) plants is believed to result from enzymatic release of HCN from the ingested organic cyanide compound. Hydrocyanic acid is readily absorbed by the guts of birds and mammals [10]. Secondary poisoning3 of terrestrial vertebrates from feeding on cyanide-poisoned invertebrates and fish is unlikely, as free cyanide is neither bioaccumulated nor persistent in the environment [1,6,10]. Because secondary poisoning is unlikely, reported anthropogenic cyanide poisonings of wildlife are usually acute events resulting from water exposure.

15.3 MECHANISMS OF TOXICITY Toxicity in animals results from the binding of cyanide to the ferric heme form of cytochrome c oxidase, which is the terminal oxidase in the mitochondrial respiratory chain [6]. This blocks electron transfer from cytochrome c oxidase to molecular oxygen, thereby inhibiting cellular respiration. This results in cellular hypoxia even in the presence of normal, oxygenated hemoglobin [6]. Hypoxia concomitantly causes a shift from aerobic to anaerobic metabolism, resulting in lactate acidosis that lowers blood pH, and depresses the central nervous system, leading to respiratory arrest and death [6]. In vivo, the majority of cyanide not complexed with heme iron can be detoxified by combining with thiosulfate to produce thiocyanate, which is excreted in the urine over a period of several days [6]. More minor detoxification pathways include exhalation of HCN and conjugation with cystene or hydroxocobalamin (vitamin B12 ) [6]. Cyanide is readily absorbed into the bloodstream and binds to hemoglobin forming methemoglobin, which is considered one of the better indicators of cytotoxicity [6].

15.4 LITERATURE REVIEW METHODS AND SCOPE Studies on cyanide toxicity to animals were obtained using both literature databases and Internet search strategies. The terms (wildlife, bird∗ , avian, shorebird∗ , waterfowl, amphibian∗ , “marine mammal,” “marine mammals”) and (toxic∗ , ecotoxic∗ , sensit∗ ) and (cyanid∗ , metallocyanid∗ , organocyanid∗ ) were used to search literature databases: ASFA, BIOSIS, CC Search® 7 Editions, Water Resources Abstracts, and Zoological Record in January–February 2003. Various search engines were used to scan the Internet for relevant articles using the keywords and phrases. These searches returned 224 records. Records were retrieved and abstracts or titles screened to judge relevance and utility, yielding 49 records. Of these, 24 were available and reviewed, and 10 were accepted as adequate studies according to the criteria described in the following section. No data for marine aquatic-dependent wildlife were found. 3 Secondary

poisoning represents toxicity to organisms that consume a cyanide-containing plant or animal.

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15.4.1 DATA QUALITY DETERMINATION Reported data were screened according to the following criteria. In some instances, these criteria could not be applied, and in some instances where data were accepted, qualifications were identified. • • • •

Primary publications were used when possible, rather than review papers. The complete study design had to be detailed in the paper. Multiple doses had to be tested with evidence of a satisfactory dose–response relationship. Studies had to report either a lethal dose for 50% of a population (LD50), or no observable adverse effect level (NOAEL) calculated using an acceptable statistical method for each endpoint measured.

15.4.2 DATA NORMALIZATION Data were normalized from the units reported in the original study to dose in units of milligrams [mg] of cyanide ion [CN] per kilogram [kg] body weight [BW] to facilitate comparison between studies. The calculations performed are outlined below.

15.4.2.1 Normalization to mg/kg Body Weight The concentration or doses of cyanide compound (CC) were converted to a standard dose of mg tested compound (TC) per kg BW using the following equations: • Dietary food concentration reported in ppm: ppm (mg CC/kg food) × average food consumption (kg food/day) = mg TC/kg BW/day average body weight (kg BW) (15.1) • Drinking water or injection concentration reported in mmol/kg: mmol CC/kg × (1 mol/1000 mmol) × molec. wt.(g/mol) × 1000 mg/g = mg TC/kg BW

(15.2)

In some subchronic and chronic tests, the doses tested changed during the study, requiring an assumption about the average dose tested. For example, one study commenced with one-day-old chicks and lasted for nine weeks, during which time the concentration of cyanide in the food remained unchanged, but the ration consumed and, hence, dose changed with time [12]. Sample et al. [13] proposed a solution for this situation in their derivation of widely-used toxicological benchmarks for wildlife. They proposed using the animal’s average body weight for the test period to calculate average food consumption using an accepted allometric equation from USEPA [14]: food consumption rate (g/day) = 0.648(BW [g])0.651

(15.3)

These values were expressed as kg food/day by multiplying by 0.001 g/kg. Sample et al. [13] note that this method over- and under-estimates food consumption (and hence dose) for younger and older chicks, respectively, but is an acceptable estimate of the average dose.

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TABLE 15.1 Cyanide Compounds Tested and Percentage Cyanide Contents Assumed in Normalizing Doses to CN Compound name

Formula

Formula weight

CN molecular weight

Acetone cyanohydrin Acetonitrile Acrylonitrile CN of cassava Hydrocyanic acid Malononitrile n-butyronitrile Potassium cyanide Propionitrile Sodium cyanide Succinonitrile

C4 H7 NO C2 H3 N C3 H3 N NA CHN C3 H2 N2 C4 H7 N KCN C3 H5 N NaCN C4 H4 N2

85.12 41.06 53.07 NA 27.03 66.07 69.12 65.12 56.10 49.01 80.10

26.02 26.02 26.02 26.02 26.02 26.02 26.02 26.02 26.02 26.02 26.02

Percent CN 30.57 63.37 49.03 100 96.26 39.38 37.64 39.96 46.38 53.09 32.48

NA = not available.

15.4.2.2 Normalization of Toxic Dose to Cyanide Ion After normalizing dosages based on total chemical concentrations, the data were normalized for CN dose (mg CN/kg BW) by accounting for the percentage of cyanide in the test compound (Table 15.1). Dosages normalized in these two manners were then compared. 15.4.2.3 Endpoint Normalization The objective of this analysis was to express all test results in terms of NOAEL values normalized to mg CN/kg BW. However, different studies reported toxicities in various ways, often hindering comparison. The methodology used by the European Commission [15] was adopted; it estimates NOAEL values by applying an uncertainty factor of 10 to the lowest observable adverse effect level (LOAEL) for a chronic endpoint, and an uncertainty factor of 100 for an LD50. These assessment factors are not well researched and are thus uncertain, especially for fast-acting gases like the free and simple cyanide compounds, which appear to possess a single mode of action.

15.5 DATA ANALYSIS Results are expressed as cumulative frequency distributions, which allowed interpretation of the data in terms of: • • • •

Relative sensitivity of birds versus mammals Relative toxicity of exposure routes No-effect levels protecting each organism group and exposure pathway Data variability

Normalized data for mammals exposed via drinking water (DW), food, and injection pathways are shown in Figure 15.1. Raw data are provided in Tables 15.2 to 15.4. Data for birds are shown with the mammalian data in Figure 15.2, and raw data are listed in Tables 15.5 to 15.7.

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100 90 Cumulative percent

80 70 60 50 40 30 20 10 0 0.001

0.01

0.1

1

10

100

mg CN/kg BW

FIGURE 15.1 Toxicity of cyanide to mammals as a function of exposure pathway, with endpoints normalized to NOAELs expressed as mg CN/kg BW (see text for details regarding normalization). Data are plotted using a cumulative distribution function of the ranked NOAELs. Data points corresponding to specific cyanide exposure pathways are denoted by (♦) for drinking water, ( ) for food, and (
) for direct injection.

Normalized cyanide NOAELs ranged from 0.005 to 80 mg CN/kg BW. The lowest estimated no-effect levels and, hence, the most sensitive endpoints were injection studies with mammals (Figure 15.1). The latter exhibited no-effect concentrations ranging from 0.005 to 1.4 mg CN/kg BW, although the majority (approximately 85%) was below 0.1 mg CN/kg BW (Table 15.4). Data representing injection studies with complex cyanides fell into the upper portion of the dataset; although the lowest NOAEL for a complex cyanide was 0.027 mg CN/kg BW, the majority of NOAELs for complex cyanides were greater than 0.13 mg CN/kg BW (Table 15.4 and Figure 15.1). Only one avian injection study was found with a NOAEL of 0.16 mg CN/kg BW based on mortality, which ranks it within the upper range of estimated mammalian NOAELs (Table 15.7 and Figure 15.2). Dietary exposures of complex cyanides appeared much less toxic to birds and mammals than those with simple cyanides. The estimated NOAELs for complex cyanides introduced via the diet ranged from 5.9 to 79.6 mg CN/kg BW (Table 15.3 and Figure 15.2), and included the single bird food ingestion study with cassava (Table 15.6) along with all of the mammalian food ingestion studies. The remaining avian food ingestion studies used sodium cyanide, which exhibited much greater toxicity with estimated NOAELs ranging from 0.014 to 0.11 mg CN/kg BW (Table 15.6). NOAELs estimated from drinking water studies for both birds and mammals fell within the ranges of most other NOAELs except for mammalian food ingestion studies (Figure 15.2). Estimated NOAELs for mammals ranged from 0.02 to 4.3 mg CN/kg BW (Table 15.2), and those for birds ranged from 0.01 to 0.8 mg CN/kg BW (Table 15.5).

15.6 DISCUSSION Data presented in the previous section will be discussed first in terms of the influence of exposure route on the relative toxicity of cyanide to mammals and birds, and then in terms of differences between mammals and birds for each exposure route.

15.6.1 ROUTE OF EXPOSURE — MAMMALS Two studies examined effects of drinking water exposure to two different wildlife species from three simple cyanide compounds [16,17]. Ballantyne’s [17] study with rabbits calculated very similar

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Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ 7 to 15 kg Unknown/ 7 to 15 kg

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Canis latrans (Coyote)

Canis latrans (Coyote)

[17]

[17]

[17]

[17]

[17]

[16]

[16]

Species

[17]

Reference

Age/ body weight Water gavage

Water gavage

Water gavage

Water gavage

Water gavage

Water gavage

Water gavage Water gavage

HCN

NaCN

KCN

NaCN

HCN

NaCN NaCN

Exposure type

KCN

Cyanide compound tested

TABLE 15.2 Toxicity of Cyanide to Mammals Exposed via Drinking Water

Effect

0.156 mmol/kg

0.117 mmol/kg

0.115 mmol/kg

0.104 mmol/kg

0.092 mmol/kg

0.09 mmol/kg

Mortality LOAEL 8 mg/kg

Mortality NOAEL 4 mg/kg

Mortality LD50

Mortality LD50

Mortality LD50

Mortality LD50

Mortality LD50

Mortality LD50

Endpoint

8

4

0.042

0.057

0.075

0.051

0.025

0.059

4.247

2.124

0.041

0.030

0.030

0.027

0.024

0.023

Concentration normalized to Concentration normalized to NOAEL mg NOAEL mg Concentration compound/kg CN/kg BW BW as reported

Study design not described in detail Study design not described in detail Study design not described in detail Study design not described in detail Study design not described in detail Study design not described in detail

Comments

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Weaning/ 87 g

Unknown 16.1 kg Unknown/ 16.1 kg Weaning/ 87 g

Unknown/ unknown

Weaning/ 87 g

Cricetomys gambianus Waterhouse Rat (African giant)

Sus sp. (Pig)

Sus sp. (Pig)

Cricetomys gambianus Waterhouse Rat (African giant)

Canis familiaris (Dog)

Cricetomys gambianus Waterhouse Rat (African giant)

[18]

[18]

[19]

[32]

[19]

Species

[19]

Reference

Age/ body weight

HCN

NaCN

CN of cassava CN of cassava HCN

HCN

Cyanide compound tested

Dietary (concentration in cassava parts)

Dietary (concentration in cassava parts) Dietary

Dietary

Dietary (concentration in cassava parts) Dietary

Exposure type

TABLE 15.3 Toxicity of Cyanide to Mammals Exposed via Food Ingestion

Unbounded NOAEL Unbounded NOAEL Unbounded NOAEL

NOAEL

Effect

NOAEL Food consumption, blood chemistry, behavior, or organ histology Growth rate LOAEL

Daily weight gain Mortality

Mortality

Growth rate

Endpoint

150 mg/kg* food

150 mg/kg

8

150

33

17.25

400 ppm 597 mg/kg* food

17.25

6

400 ppm

110 mg/kg* food

As cited by [6]

HCN measured in cassava tuber

8.068

HCN measured in cassava peel

HCN measured in cassava tuber

Comments

79.635

32.236

17.250

17.250

5.917

Concentration Concentration normalized to normalized to NOAEL mg Concentration compound/kg NOAEL mg CN/kg BW BW as reported

292 Cyanide in Water and Soil

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Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown

Oryctolagus cuniculus (Rabbit) Oryctolagus cuniculus (Rabbit) Oryctolagus cuniculus (Rabbit) Oryctolagus cuniculus (Rabbit) Felis catus (Cat)

Rattus norvegicus (Rat)

Oryctolagus cuniculus (Rabbit)

[17]

[21]

[17]

[21]

[17]

[17]

[21]

Unknown/ unknown

Oryctolagus cuniculus (Rabbit)

Species

[17]

Reference

Age/ body weight

Intravenous IV injection Intravenous Intravenous IV injection IV injection Intramuscular

HCN HCN NaCN KCN HCN HCN NaCN

Mortality LD50 0.033 mmol/kg

Mortality LD50 0.81 mg/kg

Mortality LD50 0.81 mg/kg

Mortality LD50 0.029 mmol/kg

Mortality LD50 0.025 mmol/kg

Mortality LD50 0.66 mg/kg

Mortality LD50 0.022 mmol/kg

Mortality LD50 0.018 mmol/kg

Intramuscular

HCN

Concentration as reported

Endpoint Effect

Cyanide compound Exposure type tested

TABLE 15.4 Toxicity of Cyanide to Mammals Exposed via Direct Injection

0.016

0.008

0.008

0.019

0.012

0.007

0.006

0.005

Concentration normalized to NOAEL mg compound/kg BW

0.009

0.008

0.008

0.008

0.007

0.006

0.006

0.005

Concentration normalized to NOAEL mg CN/kg BW

Male value used, female not significantly different. Study design not described in detail

Study design not described in detail Study design not described in detail

Female value used, male not significantly different. Study design not described in detail Study design not described in detail

Comments

Toxicity of Cyanide to Aquatic-Dependent Wildlife 293

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Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown

Oryctolagus cuniculus (Rabbit) Oryctolagus cuniculus (Rabbit)

Macaca mulatta (Monkey)

Canis familiaris (Dog)

Cavia porcellus (Guinea Pig)

Oryctolagus cuniculus (Rabbit) Oryctolagus cuniculus (Rabbit)

[17]

[21]

[21]

[21]

[17]

[17]

[17]

Unknown/ unknown Unknown/ unknown Unknown/ unknown

Mus musculus (Mouse)

Species

Age/ body weight

[21]

Reference

TABLE 15.4 Continued

Exposure Type IV injection Transocular injection Intramuscular

IV injection IV injection IV injection IP injection IP injection

Cyanide compound tested HCN HCN KCN

HCN HCN HCN KCN NaCN

Mortality LD50 0.057 mmol/kg

Mortality LD50 0.055 mmol/kg

Mortality LD50 1.43 mg/kg

Mortality LD50 1.34 mg/kg

Mortality LD50 1.3 mg/kg

0.028

0.036

0.014

0.013

0.013

0.031

0.010

Mortality LD50 0.039 mmol/kg Mortality LD50 0.047 mmol/kg

0.010

Concentration as reported

Mortality LD50 0.99 mg/kg

Endpoint Effect

Concentration normalized to NOAEL mg compound/kg BW

0.015

0.014

0.014

0.013

0.013

0.012

0.010

0.010

Concentration normalized to NOAEL mg CN/kg BW

Study design not described in detail Female value used, male not significantly different. Study design not described in detail

Study design not described in detail Male value used, female not significantly different. Study design not described in detail

Comments

294 Cyanide in Water and Soil

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Mortality LD50 0.093 mmol/kg 0.046

Mortality LD50 0.1 mmol/kg

Mortality LD50 8.7 mg/kg

IP injection

IP injection

IP injection

IP injection

IP injection

IP injection

NaCN

HCN

KCN

KCN

Acetone cyanohydrin

Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ 30 g

Rattus norvegicus (Rat)

Cavia porcellus (Guinea Pig)

Mus musculus (Mouse)

Cavia porcellus (Guinea Pig)

Mus musculus (CD-1 mouse)

[17]

[17]

[17]

[17]

[22]

[17]

[33]

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0.065

Mortality LD50 0.099 mmol/kg 0.065

Mortality LD50 0.098 mmol/kg 0.027

Mortality LD50 0.096 mmol/kg 0.047

Mortality LD50 0.085 mmol/kg 0.055

Mortality LD50 0.088 mmol/kg 0.057

Unknown/ unknown

Rattus norvegicus (Rat)

[17]

IP injection

Unknown/ unknown

Rattus norvegicus (Rat)

[17]

Unknown/ 250 KCN to 260 g NaCN Unknown/ unknown

Mortality LD50 0.083 mmol/kg 0.022

Mortality LD50 0.064 mmol/kg 0.017

Rattus norvegicus (Rat) (Sprague–Dawley) Mus musculus (Mouse)

IP injection

IP injection

HCN

KCN

Unknown/ unknown

IP injection

Oryctolagus cuniculus (Rabbit)

HCN

[17]

0.027

0.026

0.026

0.025

0.025

0.024

0.023

0.022

0.022

0.017

Study design not described in detail Study design not described in detail Study design not described in detail Study design not described in detail Study design not described in detail

Male value used, female not significantly different. Study design not described in detail Male value used, female not significantly different. Study design not described in detail Study design not described in detail

Toxicity of Cyanide to Aquatic-Dependent Wildlife 295

Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ unknown Unknown/ 30 g Unknown/ unknown Unknown/ unknown

Mus musculus (Mouse)

Oryctolagus cuniculus (Rabbit)

Canis familiaris (Dog)

Cavia porcellus (Guinea Pig)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

Mus musculus (CD-1 mouse)

Oryctolagus cuniculus (Rabbit)

Oryctolagus cuniculus (Rabbit)

[17]

[34]

[17]

[17]

[17]

[22]

[17]

[17]

Species

Age/ body weight

[17]

Reference

TABLE 15.4 Continued

0.112 mmol/kg

0.121 mmol/kg

0.298 mmol/kg

0.343 mmol/kg

Mortality LD50

Transocular injection Percutaneous Mortality LD50 injection IP injection

Percutaneous Mortality LD50 injection Percutaneous Mortality LD50 injection

KCN

HCN

Malononitrile NaCN

KCN

NaCN

Subcutaneous Mortality LD50 injection IP injection Mortality LD50

NaCN

Mortality LD50

18 mg/kg

0.26 mmol/kg

5.36 mg/kg

0.103 mmol/kg

Mortality LD50

Transocular injection

NaCN

0.103 mmol/kg

Mortality LD50

IP injection

Concentration as reported

HCN

Endpoint Effect

Exposure type

Cyanide compound tested

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0.146

0.18

0.071

0.079

0.055

0.054

0.050

0.028

Concentration normalized to NOAEL mg compound/kg BW

0.089

0.078

0.071

0.068

0.031

0.029

0.028

0.027

0.027

Concentration normalized to NOAEL mg CN/kg BW

Study design not described in detail Study design not described in detail

Study design not described in detail Study design not described in detail As reported by [16] Study design not described in detail Study design not described in detail Study design not described in detail

Comments

296 Cyanide in Water and Soil

Mus musculus (CD-1 mouse)

Mus musculus (CD-1 mouse)

Mus musculus (CD-1 mouse)

Rattus norvegicus (Rat) (Wistar) Mus musculus (CD-1 mouse)

Rattus norvegicus (Rat) (Wistar)

[22]

[22]

[22]

[20]

[20]

[22]

Mus musculus (CD-1 mouse)

[22] Unknown/ 30 g Unknown/ 30 g Unknown/ 30 g Unknown/ 30 g Adult/ 275 to 325 g Unknown/ 30 g Adult/ 275 to 325 g Mortality LD50 Mortality LD50

n-Butyronitrile IP injection IP injection IP injection Injection IP injection Injection

Succinonitrile Acrylonitrile KCN Acetonitrile KCN

46 mg/kg

62 mg/kg

38 mg/kg

28 mg/kg

175 mg/kg

Mortality LOAEL 3.5 mg/kg BW

Mortality LD50

Mortality NOAEL 2.5 mg/kg BW

Mortality LD50

Mortality LD50

IP injection

Propionitrile

3.5

1.75

2.5

0.46

0.62

0.38

0.28

1.398

1.109

0.999

0.226

0.201

0.143

0.130

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298

100 90 Cumulative percent

80 70 60 50 40 30 20 10 0 0.001

0.01

0.1

1

10

100

mg CN/kg BW

FIGURE 15.2 Toxicity of cyanide to birds and mammals as a function of exposure pathway, with endpoints normalized to NOAELs expressed as mg CN/kg BW (see text for details regarding normalization). Data are plotted using a cumulative distribution function of the ranked NOAELs. Data points corresponding to specific organisms and cyanide exposure pathways are denoted by () for mammals by drinking water, ( ) for mammals by food, (
) for mammals by direct injection, () for birds by drinking water, (•) for birds by food, and () for birds by direct injection.

LD50s for drinking water exposure of HCN, NaCN, and KCN: estimated NOAELs ranged from 0.023 to 0.041 mg CN/kg BW (Table 15.2). Sterner’s [16] study with coyotes observed a NOAEL and LOAEL for mortality (2.1 and 4.2 mg CN/ kg BW, respectively), that overlapped the LD50 range for rabbits (LD50s of 2.3 to 4.1 mg CN/kg BW). However, the coyote NOAEL was 100 times higher than the rabbit NOAELs (Table 15.2). None of the three studies that examined cyanide toxicity via ingestion [6,18,19] indicated that complex cyanides in food were as toxic as simple cyanides in drinking water or administered by injection. Tewe and Pessu [18] fed pigs a diet of cassava for 72 days that contained up to 400 ppm cyanide, including both free and bound cyanide with additional KCN added. Because the pigs grew significantly during the study period, the methods of Sample et al. [13] were used to estimate the average dose tested. The average weight of the pigs was 32 kg, and the food consumption given in the paper (i.e., 1.38 kg food/day) was used to calculate a dose of 17.25 mg CN/kg BW (Table 15.3). No significant mortality or weight gain differences were observed at these highest test concentrations compared to the control animal responses. Tewe’s [19] study, with African giant rats fed various parts of cassava containing differing cyanide concentrations (0, 110, 150, 597 mg HCN/kg cassava) for 16 weeks, disclosed no mortality at the highest concentration (597 mg HCN/kg cassava, or 32 mg CN/kg BW; Table 15.3), but reduced growth rate was observed at the mid and highest concentration (NOAEL of 5.9 mg CN/kg BW and LOAEL of 8.1 mg CN/kg BW). There was a discrepancy in the paper regarding the cyanide concentration units; the tables in the text reported mg/g, but the abstract and Eisler [6] reported them in mg/kg. As mg/kg was used in another Tewe study of cyanide concentrations in cassava [18], we assumed mg/kg to be the correct units. One other study reported by Eisler [6] investigated the effects of 30 days cyanide food ingestion by dogs. No sublethal effects on food ingestion were observed at 80 mg CN/kg BW, which was assumed to represent the NOAEL (Table 15.3). The greatest number of studies and endpoints of cyanide toxicity to mammals represented LD50s arising from injected doses. Ballantyne [17] studied four species (rat, mouse, rabbit, guinea pig), five different injection routes (intravenous, intramuscular, intraperitoneal, percutaneous, and transocular), and three simple cyanide compounds (HCN, NaCN, and KCN). This study observed LD50s ranging from 0.5 to 9 mg CN/kg BW (estimated NOAELs of 0.005 to 0.089 mg CN/kg BW; Table 15.4).

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Water gavage Water gavage

KCN Young adult females/1000 g KCN Young adult females/ 1000 g

Anas platyrhynchos (Mallard) Anas platyrhynchos (Mallard)

[23]

[23]

Water gavage

KCN Young adult females/1000 g

Exposure type

Anas platyrhynchos (Mallard)

Species

Cyanide compound tested

[23]

Reference

Age/ body weight

TABLE 15.5 Toxicity of Cyanide to Birds Exposed via Drinking Water

Heart ATP levels

Unbounded LOAEL

Liver and brain ATP levels Mortality

Unbounded NOAEL Unbounded NOAEL

Effect

Endpoint

0.025

2 2

0.25 mg/kg

2 mg/kg mg/kg

0.799

0.799

0.010

Concentration Concentration normalized to normalized to NOAEL mg Concentration compound/kg NOAEL mg CN/kg BW BW as reported

Toxicity of Cyanide to Aquatic-Dependent Wildlife 299

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Gelatin capsule

Gelatin capsule

NaCN

Unknown/ 185 g

Reproductively NaCN active/130 g

Otus asio (Eastern screechowl)

Coturnix japonica (Japanese quail)

[24]

Gelatin capsule

NaCN

[24]

Gelatin capsule

NaCN

Unknown/ 2215 g

[24]

Coragyps atratus (Black vulture)

Gelatin capsule

NaCN

[24]

Exposure type

Anas platyrhynchos 6-month old/ (Mallard) 1260 g Falco sparverius Unknown/ (American kestrel) 118 g

Species

Cyanide compound tested

[25]

Reference

Age/ body weight

TABLE 15.6 Toxicity of Cyanide to Birds Exposed via Food Ingestion

Mortality LD50

Mortality LD50

Mortality LD50

Mortality LD50

Mortality LD50

Endpoint Effect

9.4 mg/kg

8.6 mg/kg

4.8 mg/kg

4 mg/kg

2.7 mg/kg

0.094

0.086

0.048

0.04

0.027

Concentration normalized to NOAEL mg Concentration compound/kg BW as reported

0.050

0.046

0.025

0.021

0.014

Concentration normalized to NOAEL mg CN/kg BW

Good study with different species, no control birds reported Good study with different species, no control birds reported Good study with different species, no control birds reported Good study with different species, no control birds reported, value used with male and female data combined

Comments

300 Cyanide in Water and Soil

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Unknown/ 1610 g

1 day/ 32 g

Gallus domesticus (Chick (Broiler))

[12]

Unknown/ 75 g

Gallus domesticus (Chicken)

Sturnus vulgaris (Starling)

[24]

[24]

Gelatin capsule

Gelatin capsule

Mortality LD50

Mortality LD50

21 mg/kg

17 mg/kg

Mortality Unbounded 83 ppm CN of cassava Dietary NOAEL (concentration in cassava added to food)

NaCN

NaCN

6.024

0.21

0.17

6.024

0.111

0.090

Good study with different species, no control birds reported, value used with male and female data combined Good study with different species, no control birds reported Cassava added to food, difficult to calculate dose. Average BW used in EPA food consumption equation for birds = 38.47 g food/day

Toxicity of Cyanide to Aquatic-Dependent Wildlife

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301

[26]

Reference

Columba livia (Rock Dove)

Species

Age/ body weight KCN

Cyanide compound tested IV or intramuscular injection

Exposure type

TABLE 15.7 Toxicity of Cyanide to Birds Exposed via Direct Injection

Effect LOAEL

Endpoint Mortality

4 mg/kg

Concentration as reported

Comments As reported by [24]

Concentration normalized to NOAEL mg CN/kg BW 0.160

Concentration normalized to NOAEL mg compound/kg BW 0.4

302 Cyanide in Water and Soil

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Our review suggested that cyanide’s toxicity is remarkably similar when simple cyanide compounds were injected. Most of the estimated NOAELs (83%) ranged between 0.005 and 0.1 mg CN/kg BW (Figure 15.1). The rest ranged between 0.1 and 1.1 mg CN/kg BW (Figure 15.1, Table 15.4). For example, MacMillan [20] injected KCN into mice and observed a NOAEL and LOAEL for lethality to be 1 and 1.4 mg CN/kg BW, respectively. They also measured various energy metabolites in the brain and found that all returned to control levels within 6 to 24 h after the initial adverse effect. McNamara [21] injected HCN into seven different mammals and derived LD50s ranging from 0.6 to 1.4 mg CN/kg BW (estimated NOAELs of 0.006 and 0.014 mg CN/kg BW). Complex organic cyanides are less toxic than simple cyanides when injected into mammals. Willhite and Smith [22] investigated toxicity to rats of seven aliphatic nitriles (organic chemicals with the general formula of R–CN). Five of the seven compounds had the least toxic LD50s of all injection studies (13 to 110 mg CN/kg BW, estimated NOAELs of 0.13 to 1.1 mg CN/kg BW) (Figure 15.1, Table 15.4). There was one exception: the estimated NOAEL for acetone cyanohydrin (0.027 mg CN/kg BW) fell within the upper range of the mammalian injection data. The injection study endpoints (mostly estimated from LD50s) overlap the drinking water NOAELs, while food consumption NOAELs are greater (i.e., less toxic) than the other routes of exposure (Figure 15.1). A trend of increased tolerance, increased metabolism, and reduced bioavailability can be seen proceeding from the injection to water ingestion to dietary exposure pathways. However, this generalization should be viewed with caution owing to the paucity of food ingestion and drinking water data compared to injection data.

15.6.2 ROUTE OF EXPOSURE — BIRDS The avian drinking water data were obtained from a single study, and the results fall within the toxicity range reported for mammals exposed via drinking water. Young adult female mallards were dosed with a single gavage of potassium cyanide, and mortality and ATP concentrations were measured in the liver, brain, and heart [23]. A significant decrease in liver and brain ATP concentrations was observed at the lowest dose (0.1 mg CN/kg BW, estimated NOAEL of 0.01 mg CN/kg BW), with no significant effect on heart ATP concentrations or mortality at the highest dose (0.8 mg CN/kg BW) (Table 15.5). All ATP concentrations returned to normal by 24 h postexposure. Three studies investigated avian toxicity via food ingestion, and most NOAELs were considerably lower than those reported for mammals exposed via diet (Figure 15.2). These included two acute studies with single gelatin capsules given to the birds to derive LD50 values [24,25], and one 9-week study with cassava added to food [12]. The study by Henny et al. [25] used 6-week old mallards, while the age of the birds in the Wiemeyer et al. [24] study was not reported. However, the LD50 values were within an order of magnitude (0.027 to 0.21 mg CN/kg BW, estimated NOAELs from 0.014 to 0.11 mg CN/kg BW) for the seven different species when only simple cyanide exposures were considered (Figure 15.2, Table 15.6). In contrast, dietary cyanide was much less toxic to birds when fed complex cyanides in the form of cassava. This long-term study by Gomez et al. [12] started with one-day-old chicks fed cassava-amended feed for nine weeks. They observed no mortality with feed containing the highest dose of 83 mg CN/kg food. However, because body weight and food consumption were not reported daily and changed over time, the dose was estimated by the methods described in Section 15.4.2.1. Accordingly, the four-week body weight was calculated to be 530 g and food consumption was calculated to be 38.5 g food/day. The latter value might be slightly less than the birds were actually consuming, as the average consumption for the first four weeks was reported in the paper as 36.8 g/day and 101 g/day for the next five weeks. Because this unbounded4 NOAEL (6.02 mg CN/kg BW) falls in the range of the LD50s for the other birds (Figure 15.2, Table 15.6), 4 Unbounded refers to NOAELs that are not accompanied by an estimate of the corresponding LOAEL, or LOAELs

unaccompanied by the NOAEL. The true values of NOAELs estimated using unbounded LOAELs are more uncertain than bounded values.

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it can be inferred that the cyanide found in the cassava remains bound in the gastrointestinal tract and has limited bioavailability. This is consistent with mammalian data as discussed in the next section. To illustrate the variability inherent in the exposure equations, the use of the lowest body weight and calculated food ingestion rate produced an unbounded NOAEL of 16 mg CN/kg BW (versus 6.02 based on the average). The only studies available concerning birds dosed via direct injection were not available for direct review, and so we relied on secondary citations. One study [26] reported by Wiemeyer et al. [24] identified the minimum lethal dose of KCN injected into rock doves as being 4 mg/kg (1.6 mg CN/kg BW). This was assumed to be a LOAEL, which was divided by 10 to estimate the NOAEL of 0.16 mg CN/kg BW (Table 15.7). Overall, the estimated NOAELs for birds exposed via all exposure pathways ranged from 0.01 to 6 mg CN/kg BW (Figure 15.2). Toxicity data for all exposure routes were within the same range, with the lowest being a NOAEL for drinking water reporting no effect on liver and brain ATP concentrations. A consistent trend in the data is the small range of LD50 values reported in the Wiemeyer et al. [24] study of different bird species exposed via diet to sodium cyanide. These LD50s ranged from 2.7 to 21 mg NaCN/kg BW (estimated NOAELs of 0.014 to 0.11 mg CN/kg BW).

15.6.3 COMPARISONS BETWEEN BIRDS AND MAMMALS 15.6.3.1 Drinking Water Differences in sensitivity between birds and mammals exposed to cyanide in drinking water are difficult to ascertain from the data analyzed here, as all bird NOAELs and LOAELs were unbounded and, hence, more uncertain than bounded values. Also, the majority of mammalian data was estimated from LD50s and, hence, used the uncertain 100-fold assessment factor. Nevertheless, these data suggest that sublethal effects to birds and mammals occur at similar cyanide concentrations. For mammals, the majority of the estimated NOAELs ranged from 0.02 to 0.04 mg CN/kg BW and for birds from 0.01 to 0.8 mg CN/kg BW (Figure 15.2). 15.6.3.2 Food Greater differences were observed between birds and mammals in terms of dietary cyanide toxicity, with birds possibly being more sensitive. Avian NOAELs for simple cyanides ranged from 0.014 to 0.11 mg CN/kg BW, and those for mammals were higher (Figure 15.2). In one case, however, sensitivities appeared similar. A NOAEL from the single chronic avian study, which used feed amended with cassava (6.0 mg CN/kg BW for mortality based on a calculated exposure using median weight and food ingestion; Table 15.6) was similar to the lowest NOAEL for mammals fed cassava (5.9 mg CN/kg BW; Table 15.3). 15.6.3.3 Direct Injection Because only one bird injection study was identified, few conclusions can be drawn concerning the relative sensitivity of birds and mammals to cyanide via injection. The single avian study fell within the least sensitive 10 to 20% of mammal tests reported (Figure 15.2).

15.6.4 SIMPLE CYANIDE VERSUS COMPLEX CYANIDE COMPOUNDS Compared to injection studies with simple cyanide compounds (HCN, KCN, NaCN), the majority of the estimated mammalian NOAELs for complex cyanides were 5 to 100 times higher. The study with complex cyanides used aliphatic nitriles [22]. Presumably, cyanide in these compounds may be bound more tightly and, hence, dissociate less readily to HCN than the simple cyanide compounds. Nitrile complexes are known to be comparatively innocuous in the environment, low in chemical reactivity, and biodegradable [6].

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In contrast to the injection studies, no clear differences were apparent (Figure 15.2) between the toxicities of cyanide to birds and mammals from food amended with a simple cyanide versus complex cyanide compounds (i.e., cassava). Most of the cyanide found in cassava is part of a larger molecule that Tewe [19] and Eisler [6] refer to as cyanogenic glucoside. Its toxicity is likely due to the enzymatic release of HCN [6], but the rate of release of free cyanide from the cyanogenic glucosides has not been reported. Comparing the two bird studies, the unbounded NOAEL for mortality of birds consuming cassava for an extended period of time was about 50 times higher (i.e., less toxic) than the NOAELs for birds acutely exposed to gelatin capsules containing NaCN (Table 15.6). Although this suggests simple cyanides are more toxic, the available mammalian data do not confirm this. In the mammalian studies, the NOAEL for a study of a simple cyanide compound was more than twice that of cyanide in cassava. A possible explanation for the apparent contradiction in these results may be differential volatilization of HCN from the cassava feed owing to different methods of preparation, or different amounts of feed not immediately consumed.

15.7 BIOACCUMULATION OF CYANIDE Acute toxicity is the principal hazard posed by cyanide poisoning in wildlife. Cyanide does not bioaccumulate in animals [1] because sublethal doses are rapidly metabolized and excreted [10]. Therefore, sublethal effects are short-lived if present at all. In addition, cyanide does not appear to be persistent in the environment. Cyanides are lost from the water column due to precipitation and sedimentation, microbial degradation, and volatilization [6; and Chapter 9]. In soils, cyanides are usually complexed by trace metals, precipitated, metabolized by microorganisms, or volatilized [6; and Chapter 10]. Because cyanide does not bioaccumulate and is not persistent, risks to wildlife (e.g., eagles, osprey, mink) consuming aquatic animals exposed to cyanide are expected to be minimal (Chapter 17).

15.8 TOXICITY THRESHOLDS FOR CYANIDE Eisler’s [6] review of cyanide toxicity data suggested that free cyanide concentrations of <100 mg CN/kg diet for birds and <1000 mg CN/kg diet fresh weight for mammals would be protective of those organisms. The avian threshold was based on two dietary studies [12,27], and two drinking water studies [28,29]. The data of Gomez et al. [12] are at the high end (lesser toxicity) of the rest of the estimated doses for birds. It is possible that the cyanide in cassava used in the Gomez et al. [12] study was less bioavailable than the cyanide compounds used in the other avian food studies, resulting in an estimated safe concentration that may be too high for birds exposed to simple or free cyanide compounds. The original papers of Allen [28] and Clark and Hothem [29] were not readily available and thus were not reviewed. However, review of these works by Eisler [6] indicates that they suggested that <50 mg/l total cyanide would be safe for waterfowl when exposed via drinking water. One avian drinking water study, dosed young adult female mallards with a single gavage of potassium cyanide, and measured mortality, and liver, brain, and heart ATP concentrations [23], as discussed previously. The researchers found a significant decrease in liver and brain ATP concentrations at the lowest dose given (0.1 mg CN/kg BW, estimated NOAEL of 0.01 mg CN/kg BW), and no significant effect on heart ATP concentrations, or mortality at the highest dose (0.8 mg CN/kg BW). However, all ATP concentrations returned to normal by 24 h postexposure. This suggests that there may be short-term sublethal effects to birds at concentrations less than 50 mg/l total cyanide. Eisler’s [6] mammalian threshold of <1000 mg/kg food fresh weight is based on one study [30] with 1000 mg KCN/kg food added to the diet of weaning African giant rats (Cricetomys gambianus) for 12 weeks. This study was not included in our analysis because it examined only one dose. Eisler [6] reported that Tewe observed reduced food intake and body weight at this one concentration. Another study by Tewe [19] examined effects on giant rats fed cyanide-amended cassava. This study reported

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a NOAEL and LOAEL based on a body weight endpoint at 110 and 150 mg/kg food, significantly lower than the 1000 mg/kg food value used by Eisler [6]. No mortality was reported for cassava with HCN added at 600 mg/kg food. Therefore, the 1000 mg/kg value proposed by Eisler [6] may be too high to protect all mammals from the effects of dietary cyanide exposure. From our review, doses of less than 0.01 mg CN/kg BW would be fully protective of birds and mammals exposed to cyanide via both drinking water and food. While this is orders of magnitude more conservative than the 1000 mg/kg threshold suggested by Eisler [6], a threshold of 0.01 mg CN/kg BW includes both water and dietary exposure pathways, and is consistent with sensitivity distributions for the most toxic simple cyanides regardless of exposure pathway (Figure 15.2). Ongoing studies also suggest that this threshold — when converted to equivalent aqueous concentrations (µg CN/l) using allometric equations — is also consistent with AWQC for protection of aquatic organisms [31]. Given that all vertebrate organisms share the same mode of action for acute cyanide mortality [1,6], consistent toxicity thresholds from aqueous exposure are to be expected, and help confirm our more conservative tissue threshold (0.01 mg CN/kg BW) for protection of aquatic-dependent wildlife.

15.9 SUMMARY AND CONCLUSIONS • Estimated NOAELs for cyanide in animals range from 0.005 to 80 mg CN/kg BW for birds and mammals exposed to all cyanide compounds via food, drinking water, and injection. • Injection studies with mammals were the most sensitive with reported NOAELs ranging from 0.005 to 1.1 mg CN/kg BW; however, the majority (about 85%) was below 0.1 mg CN/kg BW. Data points from injection studies with complex cyanides were less sensitive, with the majority of values exceeding 0.13 mg CN/kg BW. • Both birds and mammals exhibit similar toxicity to cyanide when exposed via drinking water; NOAELs for mammals range from 0.02 to 2.1 mg CN/kg BW, and NOAELs for birds range from 0.01 to 0.8 mg CN/kg BW. These values overlap strongly with injectionbased NOAELs, suggesting a consistent toxicity of simple cyanides regardless of exposure pathway. • The highest (least toxic) estimated NOAELs range from 6 to 80 mg CN/kg BW, from a single study of birds fed complex cyanides contained in food (cassava), and from all of the mammal food cyanide ingestion studies with complex cyanides. Presumably, the complex cyanide compounds dissociate less to HCN than the simple cyanide compounds. • Acute toxicity appears to be the principal hazard posed by cyanide poisoning in wildlife. Cyanide does not appear to bioaccumulate in animals because sublethal doses are rapidly metabolized and excreted. Because cyanide does not bioaccumulate, effects to wildlife (e.g., eagles, osprey, mink) consuming aquatic animals exposed to cyanide are projected to be insignificant. • From our review, doses of less than 0.01 mg CN/kg BW are expected to be protective of birds and mammals exposed to cyanide via drinking water and food. While this is conservative, a threshold of 0.01 mg CN/kg BW protects from both water and dietary exposure pathways, and is consistent with sensitivity distributions for the most toxic simple cyanides regardless of exposure pathway.

ACKNOWLEDGMENTS This review was conducted in part with support from the Water Environment Research Foundation (WERF), Project #01-ECO-1. We thank the WERF project coordinator, Margaret Stewart, and the project subcommittee (Walter Berry, Phillip Dorn, Joseph Gorsuch, Jim Pletl, and Mary Reiley)

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for their review of the manuscript. We also thank Joe Volosin for assistance with graphics, and Diane Gensemer for proofreading and copyediting.

REFERENCES 1. USEPA, Ambient water quality criteria for cyanide — 1984, Report EPA 440/5-84-028, U.S. Environmental Protection Agency, Washington, DC, 1985. 2. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Report 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 3. Gensemer, R.W., DeForest, D., Coyner, A., Clark, J., Cardwell, R.D., Dzombak, D.A., Higgins, C., and Santore, R.C., Reassessment of cyanide criteria for the protection of aquatic life and wildlife: phase I annual progress report (01-ECO-1), Water Environment Research Foundation, Alexandria, VA, 2003. 4. Newman, M.C. and Jagoe, C.H., Ligands and the bioavailability of metals in aquatic environments, in Bioavailability: Physical, Chemical, and Biological Interactions, Hammelink, J.L., Landrum, P.F., Bergman, H.L., and Benson, W.H., Eds., Lewis Publishers, Boca Raton, FL, 1994, p. 39. 5. USEPA, Screening level ecological risk assessment protocol for hazardous waste combustion facilities, Report EPA 530-D-99-001A, U.S. Environmental Protection Agency, Washington, DC, 1999. 6. Eisler, R., Cyanide hazards to fish, wildlife, and invertebrates: a synoptic review, Biological Report 85(1.23), U.S. Fish and Wildlife Service, Washington, DC, 1991, p. 55. 7. Eisler, R., Clark, D.R., Wiemeyer, S.N., and Henny, C.J., Sodium cyanide hazards to fish and other wildlife from gold mining operations, in Environmental Impacts of Mining Activities: Emphasis on Mitigation and Remedial Measures, Azcue, J.M., Ed., Springer-Verlag, Berlin, 1999, p. 55. 8. Doudoroff, P., Leduc, G., and Schneider, C.R., Acute toxicity to fish of solutions containing complex metal cyanides, in relation to concentrations of molecular hydrocyanic acid, T. Am. Fish. Soc., 95, 6, 1966. 9. Doudoroff, P., Toxicity to fish of cyanides and related compounds, Report EPA-600/3-76-038, U.S. Environmental Protection Agency, Cincinnati, OH, 1976. 10. Hill, E.F. and Henry, P.F.P., Cyanide, in Noninfectious Diseases in Wildlife, Hoff, G.L., Ed., Iowa State University Press, Ames, Iowa, 1996, p. 99. 11. Padmaja, G., The culprit in cassava toxicity: cyanogens or low protein?, Consultative Group on International Agricultural Research, News Letter, 3, www.worldbank.org/html/cgiar/newsletter/Oct96/ 6cgnews.html, 1996. 12. Gomez, G., Aparico, M.A., and Willhite, C.C., Relationship between dietary cassava cyanide levels and broiler performance, Nutr. Rep. Int., 37, 63, 1988. 13. Sample, B.E., Opresko, D.M., and Suter II, G.W., Toxicological benchmarks for wildlife: 1996 revision, Report ES/ER/TM-86/R3, U.S. Department of Energy, Office of Environmental Management, Oak Ridge, TN, 1996. 14. USEPA, Wildlife exposure factors handbook, Report EPA/600/R-93/187, U.S. Environmental Protection Agency, Office of Research and Development, Washington DC, 1993. 15. European Commission, Technical guidance document in support of commission directive 93/67/EEC on risk assessment for new notified substances and commission regulation (EC) No. 1488/94 on risk assessment for existing substances, Office of Official Publication of the European Communities, Brussels, Luxembourg, 1996. 16. Sterner, W.S., Effects of sodium cyanide and diphacinone in coyotes (Canis latrans): applications as predacides in livestock toxic collars., B. Environ. Contam. Tox., 23, 211, 1979. 17. Ballantyne, B., The influence of exposure route and species on the acute lethal toxicity and tissue concentrations of cyanide, in Developments in the Science and Practice of Toxicology, Hayes, A.W., Schnell, R.C., and Miya, T.S., Eds., Elsevier, Amsterdam, 1983, p. 583. 18. Tewe, O.O. and Pessu, E., Performance and nutrient utilization in growing pigs fed cassava peel rations containing different cyanide levels, Nutr. Rep. Int., 26, 51, 1982. 19. Tewe, O.O., Effect of cassava-based diets varying in cyanide content on the performance and physiopathology of the African giant rat (Cricetomys gambianus Waterhouse), Anim. Feed Sci. Tech., 11, 1, 1984.

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20. MacMillan, V.H., Cerebral energy metabolism in cyanide encephalopathy, J. Cerebr. Blood F. Met., 9, 156, 1989. 21. McNamara, B.P., Estimates of the toxicity of hydrocyanic acid vapors in man, Report EB-TR-76023, U.S. Department of the Army, Edgewood Arsenal, Aberdeen Proving Ground, MD, 1976. 22. Willhite, C.C. and Smith, R.P., The role of cyanide liberation in the acute toxicity of aliphatic nitriles, Toxicol. Appl. Pharmacol., 59, 589, 1981. 23. Ma, J. and Pristos, C.A., Tissue-specific bioenergetic effects and increased enzymatic activities following acute sublethal peroral exposure to cyanide in the mallard duck, Toxicol. Appl. Pharm., 142, 297, 1997. 24. Wiemeyer, S.N., Hill, E.F., Carpenter, J.W., and Krynitsky, A.J., Acute oral toxicity of sodium cyanide in birds, J. Wildlife Dis., 59, 589, 1986. 25. Henny, C.J., Hallock, R.J., and Hill, E.F., Cyanide and migratory birds at gold mines in Nevada, USA, Ecotoxicology, 3, 45, 1994. 26. Spector, W.S., Acute toxicities of solids, liquids, and gasses to laboratory animals, in Handbook of Toxicology, Vol. 1, WB Saunders Company, Philadelphia, PA, 1956, p. 408. 27. Gomez, G., Valdivieso, M., Santos, J., and Hoyos, C., Evaluation of cassava root meal prepared form low- or high-cyanide containing cultivars in pig and broiler diets, Nutr. Rep. Int., 28, 693, 1983. 28. Allen, C.H., Mitigating impacts to wildlife at FMC Gold Company’s Paradise Peak mine, in Proceedings of Nevada Wildlife/Mining Workshop, Reno, NV, March 27–29, 1990. 29. Clark, D.R. and Hothem, R.L., Mammal mortality at Arizona, California, and Nevada gold mines using cyanide extraction, California Fish Game, 77, 61, 1991. 30. Tewe, O.O., Effect of dietary cyanide on the performance, metabolism and pathology of the African rat (Cricetomys gambianus Waterhouse), Nutr. Rep. Intl, 26, 529, 1982. 31. Gensemer, R.W., Volosin, J., Clark, J.M., and Cardwell, R.D., Are ambient water quality criteria for cyanide protective of aquatic-dependent wildlife?, in Proceedings of Fourth SETAC World Congress, Society of Environmental Chemistry and Toxicology, Portland, OR, USA, 2004, p. 258. 32. USEPA, Ambient water quality criteria for cyanide, Report EPA 440/5-80-037, U.S. Environmental Protection Agency, Office of Water, Washington DC, 1980. 33. Keinston, R.C., Cabellon, J.S., and Yarbrough, K.S., Pyridoxal 5 -phosphate as an antidote for cyanide, spermine, gentamicin, and dophamine toxicity: an in vivo rat study, Toxicol. Appl. Pharm., 88, 433, 1987. 34. Chen, K.K. and Rose, C.L., Nitrite and thiosulfate therapy in cyanide poisoning, JAMA, 149, 113, 1952.

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Health Risk Assessment 16 Human of Cyanide in Water and Soil Barbara D. Beck, Mara Seeley, Rajat S. Ghosh, Peter J. Drivas, and Neil S. Shifrin CONTENTS 16.1

Environmental Concentrations and Exposure Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.1.1 Concentrations of Cyanide Compounds in Soil at Former MGP Sites . . . . . . . . . 16.1.2 Concentrations of Cyanide Compounds in Groundwater at Former MGP Sites 16.1.3 Estimated Concentrations of Cyanide Compounds in Air at Former MGP Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.2 Exposure Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.2.1 Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.2.2 Ferric Ferrocyanide Solid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.2.3 Dissolved Ferrocyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3 Toxicity Evaluation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.1 Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.1.1 Subchronic/Chronic Toxicity Criterion — Inhalation . . . . . . . . . . . . . . . 16.3.1.2 Subchronic/Chronic Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . 16.3.1.3 Acute Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.2 Ferric Ferrocyanide Solid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.2.1 Subchronic/Chronic Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . 16.3.2.2 Acute Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.3 Dissolved Ferrocyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3.3.1 Subchronic/Chronic Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . 16.3.3.2 Acute Toxicity Criterion — Ingestion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4 Risk Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1 Exposure Algorithms and Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1.1 Soil. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1.2 Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1.3 Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.2 Case Study Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.3 Regulatory Criteria for Cyanide Compounds in Environmental Media . . . . . . . . 16.4.3.1 Soil. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.3.2 Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5 Recommendations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

310 310 311 312 313 314 315 315 315 316 316 317 318 318 319 319 319 320 320 320 321 321 322 322 323 324 324 325 326 326 327

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Cyanide compounds associated with past and current industrial operations may be present in soils and water. As discussed in Chapter 4 and elsewhere in this book, present operations that may release cyanide compounds to the environment include electroplating, mining, iron and steel, and chemical industries. With respect to past releases, a well-known example is provided by the approximately 1000 to 3000 former manufactured gas plants (MGPs) in the United States, facilities that made gas for cooking, lighting, and heating for over a century [1]. MGP sites today are sometimes listed on state hazardous waste site cleanup programs, with cyanide being a frequent compound of interest [2]. Evaluating potential human health risks of cyanide in water and soil is complicated by the chemical differences among cyanide compounds. Cyanide chemistry and environmental conditions influence potential human health risk in a number of ways — potential for release from an environmental source (e.g., soil) to an exposure medium (e.g., air), and uptake into the body as a function of exposure route, exposure duration, and inherent toxicity. For example, acute exposure to free cyanide (HCN, CN− ) in air and water may be an issue in occupational settings, but exposures to cyanide at inactive MGP sites are typically subchronic (weeks to months) or chronic (at least 10% of a lifespan), and involve primarily iron–cyanide compounds in water and soil. In this chapter, we discuss potential human health risks associated with cyanide compounds found in water and soil. To provide a framework to illustrate the human health risk assessment process, we focus on specific cyanide compounds found at MGP sites, as there is a wide range in toxicity of the various cyanide species found at these sites (Chapter 13). Of particular interest at MGP sites is dissolved free cyanide (HCN, CN− ), which is the most toxic form of cyanide; solid phase iron–cyanide compounds, specifically ferric ferrocyanide solid [Fe4 (Fe(CN)6 )3 (s)]; and dissolved 4− ferrocyanide resulting from the dissolution of ferric ferrocyanide solid [Fe(CN)3− 6 ; Fe(CN)6 ]. In the context of an MGP site scenario, we describe typical exposure pathways and the toxicity of the compounds (including toxicity criteria), and provide a screening level case study to estimate risks using upper bound parameters for media concentrations and receptor intakes. We then present a perspective on the results of the risk assessment, through comparison with other exposures to cyanide, followed by recommendations for improving risk assessment procedures for cyanide compounds at contaminated sites.

16.1 ENVIRONMENTAL CONCENTRATIONS AND EXPOSURE PATHWAYS This section provides estimates of concentrations of free and iron-complexed cyanide in soil, water, and air at MGP sites, to provide a basis for exposure evaluation. Some estimates are data-derived, whereas other estimates are modeled. Many of the estimates represent conservative values, that is, the concentrations that are likely to overestimate actual concentrations at many sites. Thus, this section provides the foundation for the screening risk calculations presented later in the chapter.

16.1.1 CONCENTRATIONS OF CYANIDE COMPOUNDS IN SOIL AT FORMER MGP SITES Cyanide compounds found in soils at MGP sites are usually derived from oxide box waste (OBW) materials [3]. The OBW originated from the purification step in gas production, in which sulfides and cyanides (primarily hydrogen cyanide and ammonium ferrocyanide) were typically removed from the gas stream via binding to ferric oxide on a packing material, such as wood chips. OBW material contains a high percentage of ferric ferrocyanide solid, also known as Prussian Blue, and is often characterized by a striking blue color (see Chapters 2 and 5). Previous studies have found concentrations of total cyanide in OBW ranging from 1 to 2% by weight (10,000 to 20,000 mg

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total cyanide/kg OBW) [4–6]. Because of the environmental stability of the ferric ferrocyanide material, cyanide compounds are still found today in soil at MGP sites. Studies performed at ten MGP sites in New York State (hereinafter referred to as the “New York Study”) found concentrations of total cyanide in soil ranging from 11 to 20,701 mg/kg, with solid-phase and dissolved iron–cyanide complexes constituting greater than 99% of the total cyanide [7]. A similar survey of 12 MGP sites in the Netherlands found total iron cyanide concentrations in soil ranging from 34 to 20,850 mg/kg, with a median of 632 mg/kg, as ten sites had concentrations less than 4000 mg/kg [8]. Assuming the percent of iron cyanide at the MGP sites in the Netherlands is comparable to that at the New York sites, then the total iron–cyanide concentrations measured at the Netherlands sites would closely approximate total cyanide concentrations. For the purposes of this analysis, we assumed a conservative estimate of total cyanide in soil of 5000 mg/kg. We recognize that uniform concentrations of this magnitude are unrealistic. However, this analysis is for screening purposes, that is, to determine whether a more refined analysis is required, or if risks are low enough that additional analysis is unnecessary. Even though total cyanide concentrations in soil at MGP sites can be on the order of thousands of parts per million, free cyanide is usually detected only in trace quantities. Because of the greater toxicological significance of free cyanide, it is important to estimate its potential concentrations in soil. In general, the primary health concern arising from metal–cyanide compounds relates to their ability to release free cyanide (Chapter 13). The amount of free cyanide in soil at MGP sites is generally considered to be small. Theis et al. [5] found that over 97% of total cyanides in soil at MGP sites was dominated by the ferrocyanide ion, Fe(CN)4− 6 , and no free cyanide was detected. Similarly Gould et al. [4] found less than 0.1% of total cyanide in weathered OBW was present as free cyanide. According to the New York study, free cyanide concentrations at ten MGP sites in New York were below 2.5 mg/kg (<0.4% of the total cyanide) for the majority of soil samples [7]. The assumption that 0.1% of the total cyanide measured in soil was free cyanide in a 1992 study by Meeussen et al. [8] would yield a range of 0.03 to 20.8 mg free cyanide/kg soil, with a median value of 0.6 mg/kg for the MGP sites in the Netherlands. For our screening analysis, we assumed a total cyanide concentration of 5000 mg/kg soil and a free cyanide concentration of 5 mg/kg soil (i.e., 0.1% of the total cyanide concentration).

16.1.2 CONCENTRATIONS OF CYANIDE COMPOUNDS IN GROUNDWATER AT FORMER MGP SITES Dissolution of ferric ferrocyanide solid under a range of pH and redox conditions will release 3− dissolved iron–cyanide complexes [Fe(CN)4− 6 , Fe(CN)6 ] to the groundwater at MGP sites [8–11]. Dissociation of these dissolved iron–cyanide complexes is extremely slow in the absence of light, with half lives ranging from 100 to 1000 years under typical groundwater pH and redox conditions [12,13]. The rate of dissociation is greatest under conditions of low pH, whereas higher pH levels (pH > 6) promote greater stability to the iron–cyanide complex [13]. However, trace amounts of free cyanide and weak metal–cyanide complexes, such as copper and nickel cyanide, can form in the subsurface environment, due to dissociation of the iron–cyanide complexes. There are only a few comprehensive surveys regarding concentrations of total and free cyanide in groundwater at MGP sites. The 1992 study by Meeussen et al. [13] provided a possible range of free cyanide in groundwater as 0.01 to 0.1 mg/l, while soluble iron–cyanide complexes ranged from 1.7 to 11.4 mg/l. In the New York study [7], a comprehensive characterization of groundwater samples was performed at ten different MGP sites. According to species-specific analytical methodologies, free cyanide concentrations were less than 0.03 mg/l, with strong iron–cyanide complexes constituting more than 97% of the total cyanide in groundwater samples, which ranged from 0.15 to 18.7 mg/l. For this analysis, we assumed a free cyanide concentration of 0.1 mg/l, and a soluble ferrocyanide concentration of 11 mg/l.

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16.1.3 ESTIMATED CONCENTRATIONS OF CYANIDE COMPOUNDS IN AIR AT FORMER MGP SITES Cyanide is rarely monitored in air at MGP sites. Nonetheless, because of the high toxicity of hydrogen cyanide (HCN) gas, as well as the possibility that dissolved iron–cyanide complexes in adjoining surface water could photodissociate under conditions of strong ultraviolet light and alkaline pH [13], it is relevant to estimate potential air concentrations to address inhalation exposures. For this example, we have assumed a hypothetical site with a total cyanide concentration of 5000 mg/kg in soil, distributed uniformly over a one-acre site down to a depth of one foot. As noted earlier, this is an unrealistically high concentration, but appropriate for a screening analysis. Because the air model used to estimate air concentrations requires local information on meteorology, we have arbitrarily located the hypothetical MGP site in Hartford, Connecticut. Specifics of the air modeling exercise are provided below. Estimates of emissions of HCN to air were based on the assumption that dissolution of ferric ferrocyanide solid yields dissolved ferrocyanide, which subsequently decomposes to form gaseous HCN that is then released to the air. An air model recommended by the U.S. Environmental Protection Agency (USEPA) was then used along with site-specific meteorological conditions to predict air concentrations of HCN at a one-acre site. It should be noted that for estimating air impacts, decomposition of iron–cyanide complexes in soil solutions is typically not significant. Because of the slow rate of decomposition, the quantum of release of HCN to air through this pathway is usually negligible under most circumstances [14]. Surface photolysis of ferrocyanide complex under ideal conditions (e.g., hydraulic connection between impacted groundwater and a surface water body, high sunlight intensity, absence of turbidity, alkaline pH, stagnant surface conditions) could, however, release free cyanide [15], the impacts of which could best be assessed via surface water quality modeling at the discharge location (Chapter 9). The calculations below present a worst-case analysis based on measured decomposition rates of iron–cyanide complexes in groundwater and soil solutions. Although these cyanide complexes can find their way into surface water bodies, where photodissociation can strongly affect conversion to HCN, their concentration in surface waters would be far less than in MGP site groundwater. Based on the results shown in Figure 8 of the 1992 study by Meeussen et al. (and as discussed in Section 16.1.2) the half-life of iron–cyanide complexes in groundwater or soil solutions can range from about 100 to 1000 years, depending on soil pH and redox conditions [13]. For the screening analysis, which is a worst-case scenario, we chose a half-life of 100 years, resulting in a fractional decomposition rate of approximately 0.007 per year (i.e., slightly less than 1% per year). We also assumed that all of the decomposed cyanide complex would be released to the air in the form of gaseous HCN. The fractional decomposition rate of 0.007 per year was then multiplied by the total cyanide concentration in site soil and a total volume of MGP material, yielding an estimated emission rate in terms of g/sec/m2 . As noted earlier, we assumed a total cyanide concentration of 5000 mg/kg, distributed uniformly over a one-acre site down to a depth of one foot, recognizing that this is a conservatively high concentration. The estimated emission rate in terms of total cyanide (as CN) had to be converted to HCN, by multiplying by the molecular weight ratio of HCN/CN of 27/26. This resulted in an HCN emission rate of 6.3×10−7 g/sec/m2 for the one-acre site. Once released into the air, the HCN emissions from soil are transported and dispersed by winds over the surrounding area. The predicted long-term HCN air concentrations were estimated using the current version of the USEPA Industrial Source Complex, Short Term (ISCST3) air model, Version 02035, which is recommended by USEPA for regulatory purposes [16]. The ISCST3 air model can predict the maximum short-term and average long-term air concentrations resulting from a large number of combined point, volume, and area sources. The meteorological input used for the ISCST3 air modeling consisted of hourly surface wind data taken over one year at the Hartford, Connecticut airport. This meteorological data show a predominant wind direction from the north, which would cause the highest air concentrations to occur near the

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100

0.5

0.2

0.2

80

0.5

60 40

1

0.5

0.2

0

0.5

North (m)

20

–20 –40 1

–60

5

–100 –100

0.

0.2

–80

–80

–60

–40

–20

0

20

40

60

80

100

East (m)

FIGURE 16.1 Contours of predicted HCN air concentrations (µg/m3 ) for 1-acre MGP site (shaded) containing 5000 mg/kg cyanide complexes down to 1-ft depth.

southern border of the site. For air modeling purposes, the one-acre site was assumed to be a square area source, oriented north–south as shown in Figure 16.1. Flat terrain was assumed in the vicinity of the site. Calculations were made for over 1000 receptor locations, which were at a typical breathing height of 1.5 m, distributed over a 200 m × 200 m (10-acre) modeling area, with the one-acre site centered in the 10-acre modeling region. The source area and the long-term air modeling results, in terms of contours of HCN concentrations (in µg/m3 ), are presented in Figure 16.1. As shown in this figure, the maximum predicted long-term HCN air concentration was about 1 µg/m3 , located over the southern portion of the one-acre waste site, and extending to about 20 m beyond the southern boundary of the site. Predicted HCN air concentrations beyond about 50 m from the one-acre waste site were below 0.5 µg/m3 . For purposes of this case study, we assumed that an off-site resident would be exposed to an HCN concentration of 0.5 µg/m3 .

16.2 EXPOSURE PATHWAYS Definition of typical exposure pathways is critical to estimating the risks from complex and free cyanides at former MGP sites. Tables 16.1 to 16.3 present the exposure routes and likely exposure durations (i.e., chronic or acute) for this screening analysis, involving free cyanide in air, water, and soil; ferric ferrocyanide solid in soil; and dissolved ferrocyanide in water. Note that the risk from dissolved iron–cyanide complexes (i.e., soluble ferrocyanide complexes) will be related to their potential for dissociation under site-specific environmental conditions, which is considered here.

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TABLE 16.1 Soil Pathways Evaluated for MGP Site Risk Screening Analysis Ingestion

Free cyanide Ferric ferrocyanide solid

Dermal

Acute

Subchronic/chronic

Acute

Subchronic/chronic

+ −

+ +

− −

+ +

+pathway evaluated ; −pathway not evaluated.

TABLE 16.2 Water Pathways Evaluated for MGP Site Risk Screening Analysis Ingestion

Free cyanide Dissolved ferrocyanide

Dermal

Acute

Subchronic/chronic

Acute

Subchronic/chronic

− −

+ +

− −

+ +

+pathway evaluated; −pathway not evaluated.

TABLE 16.3 Air Pathways Evaluated for MGP Site Risk Screening Analysis Inhalation

Free cyanide

Acute

Subchronic/chronic

−

+

+pathway evaluated; −pathway not evaluated.

16.2.1 FREE CYANIDE For free cyanide, we evaluated ingestion of, and dermal contact with, soil and water, for both subchronic and chronic exposure durations. We also evaluated the acute exposure pathway for ingestion of free cyanide in soil, based on the assumption that a child may exhibit pica behavior for soil, that is, a child may intentionally ingest large amounts of soil (a pica child). Pica soil ingestion, in which a child ingests a much greater amount of soil than a typical child, is very rare. The pica child assumed for the screening level analysis may experience a single or limited number of events of high soil consumption. We did not include the dermal uptake pathway for acute exposures to free cyanide in soil, given that the nature of dermal contact with soil is typically of a more chronic nature. Since the release of free cyanide into air takes place over many years, and thus, is not expected to result in short-term peak concentrations, we have evaluated only subchronic and chronic air exposures to free cyanide.

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16.2.2 FERRIC FERROCYANIDE SOLID Although dermal uptake through contact with ferric ferrocyanide in soil during site remedial activities is expected to be very low [17], we have conservatively included this pathway for subchronic and chronic exposures. Because ferric ferrocyanide solid has relatively low toxicity [18], we eliminated the acute exposure pathway as being of interest. We have also eliminated the inhalation pathway for ferric ferrocyanide. Ferric ferrocyanide is a solid at room temperature, and is unlikely to yield HCN to a significant extent [19]. Although inhalation of wind-blown particles containing ferric ferrocyanide may be a possible exposure pathway, this pathway typically results in lower intake than ingestion of, or dermal contact with, soils. Thus, the exposure pathways considered for ferric ferrocyanide solid were ingestion of, and dermal contact with, soil, for subchronic and chronic exposure durations.

16.2.3 DISSOLVED FERROCYANIDE Dissolved ferrocyanide is very stable in groundwater. Due to this stability, we have evaluated exposure to dissolved ferrocyanide via ingestion of, and dermal contact with, groundwater. While ferrocyanide dissociates very slowly in the dark, it is important to recognize that when discharged to a surface water, it can undergo photodissociation, depending on various environmental factors such as sunlight intensity, pH, temperature, turbidity, and depth to water column (discussed in Chapter 9). Recent data collected from a surface wetland treatment system in Tennessee indicates a first-order ferrocyanide dissociation rate of 2.2%/h for a flow of 10 gpm [20]. Chapter 9 provides information about first-order ferrocyanide photodissociation rates in fast-moving, opensurface streams that could be as high as 40%/h. However, it should be noted that in open stream channels, there could be aerobic biodegradation of free cyanide generated in the water column. This phenomenon is most likely to occur in slow moving surface water bodies, such as a wetland system or a shallow pond, where the biodegradation rate of free cyanide can be many orders of magnitude higher than the photodissociation rate of dissolved ferrocyanide [20]. Hence, it is expected that there will be insignificant accumulation of any free cyanide in slow-moving surface water bodies.

16.3 TOXICITY EVALUATION This section describes the toxicity of free cyanide, ferric ferrocyanide solid, and dissolved ferrocyanide, considering exposure pathways as well as durations of exposure. Additional information is provided in Chapter 13. Here, we identify mechanisms of toxicity, modifying factors, and critical studies for use in developing toxicity criteria. We then discuss toxicity criteria for use in risk assessment of free cyanide, ferric ferrocyanide solid, and dissolved ferrocyanide. Note that there are no dermal toxicity criteria for either free cyanide or ferrocyanide (solid or dissolved). In the absence of dermal-specific toxicity criteria, oral toxicity criteria are used, assuming that once a chemical is absorbed into the bloodstream, the health effects are similar, regardless of whether the route of exposure is oral or dermal. For purposes of this screening analysis, we therefore assumed that free cyanide, ferric ferrocyanide, and dissolved ferrocyanide are absorbed through the skin (from both soil and water) to the same extent as through the gastrointestinal (GI) tract (e.g., via the oral route). However, note that this assumption is likely to overestimate risks from these cyanide compounds. This is because the skin is much less permeable than the GI tract. In contrast to the GI tract, where absorption takes place across a thin, single-cell layer, the skin is composed of several layers, including an outer layer (the stratum corneum) of densely packed keratinized cells, which significantly impedes absorption of most compounds, particularly charged compounds such as the 4− free cyanide anion and dissolved Fe(CN)3− 6 and Fe(CN)6 [21].

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16.3.1 FREE CYANIDE The toxicity of free cyanide is largely a function of its ability to bind tightly to enzymes containing metal ions, most notably the mitochondrial enzyme cytochrome c oxidase. Since cytochrome c oxidase is essential to aerobic metabolism, binding of free cyanide to this enzyme is thought to be particularly critical for the toxicity of free cyanide. The brain and certain other tissues rely solely on oxidative (i.e., aerobic) metabolism, and cannot switch to anaerobic metabolism when cytochrome c oxidase is inactivated by free cyanide. Thus, free cyanide is highly toxic to the brain and the central nervous system (CNS) [22,23]. The shape of the dose–response curve for free cyanide is remarkably steep. For example, in rats exposed to acetone cyanohydrin (which is hydrolyzed rapidly to HCN under physiological conditions), there were no adverse systemic effects at acetone cyanohydrin concentrations up to 211 mg/m3 (approximately 67 mg/m3 HCN), but there was 30% mortality at an acetone cyanohydrin concentration of 225 mg/m3 (approximately 71 mg/m3 HCN) [24]. This is likely a reflection of cellular function being unimpaired as long as there is sufficient oxidative metabolism. When oxidative metabolism drops below a critical threshold, physiological function is impaired, resulting in toxicity [25]. Under certain conditions, free cyanide induces toxic effects very rapidly. This rapid action is a function of how quickly free cyanide is absorbed [23]. If exposure and absorption of free cyanide occur more slowly, the liver can detoxify cyanide more readily. Hence, a large single dose that may saturate detoxification mechanisms in the liver, will produce more toxicity than the same amount delivered in multiple smaller doses that individually do not exceed the liver’s detoxification mechanisms’ capacities. Similarly, ingestion of cyanide in food, which both slows absorption and occurs over longer time periods, can reduce cyanide toxicity several-fold [26]. The features of acute cyanide poisoning are relatively similar across exposure routes, although the actual potency may vary. Characteristic signs include rapid breathing, gasping, arrhythmia, coma, and eventual death. There are no histological changes that can be detected via microscope that are diagnostic for acute cyanide poisoning [27]. Features of chronic free cyanide toxicity include those of acute free cyanide toxicity as well as neurological changes, such as damage to the corpus callosum (a thick bundle of nerve fibers connecting to the right and left cerebral hemispheres in the brain) and the optic nerve, as observed in rats [28]. There is limited information regarding features of chronic toxicity in humans. Optic neuropathy and various other neurological effects have been observed in populations consuming large quantities of cassava, which contains relatively high concentrations of cyanide [29]. Some of the chronic changes appear to be a consequence of demyelination of nerve fibers induced by free cyanide [27]. In addition, chronic cyanide toxicity is associated with impairment of thyroid function [30,31]. This is an indirect effect, due to inhibition of iodide uptake into the thyroid gland by thiocyanate, the metabolite of free cyanide. Impaired thyroid function results in reduced levels of thyroid hormones, with possible consequences on thyroid hormone regulation of metabolism, reproduction, and development.

16.3.1.1 Subchronic/Chronic Toxicity Criterion — Inhalation This section presents toxicity criteria for evaluating air concentrations of free cyanide. It should be noted that while exceedance of a criterion does not necessarily imply that adverse effects are likely, it does imply that additional analysis is warranted. For example, the criteria discussed below are protective against lifetime exposure. If the exposure duration for a site under study is on the order of weeks to months, then exceedances of the criteria may present less (or even no) toxicological concern. Several occupational studies are available regarding the effects of subchronic and chronic inhalation exposure to free cyanide in humans. While these studies have limitations, in particular incomplete exposure data and presence of other chemicals, they nonetheless present a consistent picture of CNS and thyroidal effects, associated with longer-term exposure to free cyanide [30,32,33]. In a study of

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electroplating workers in Egypt exposed to free cyanide over 5 to 10 years, El Ghawabi and coworkers [30] reported a lowest observed adverse effect level (LOAEL) of 6.4 ppm (7.1 mg/m3 ), associated with CNS symptoms (e.g., dizziness, fatigue), enlarged thyroid, and increased radioactive iodide uptake by the thyroid. Although an increase in thyroidal radioactive iodide uptake is opposite to what was expected, the study authors hypothesized that the increase could be explained as a consequence of acute cyanate withdrawal. Specifically, the workers in the study had been away from work for the two days preceding the uptake test, possibly resulting in elevated accumulation of iodide by an iodide-deficient thyroid. Because this LOAEL is associated with an occupational exposure (i.e., 8 h/day, 5 days/week), it must be adjusted to develop an equivalent concentration for environmental exposures (i.e., 24 h/day, 7 days/week), as follows: 7.1 mg/m3 ×

10 m3 /day (8-h occupational inhalation rate) × 5 days = 2.5 mg/m3 20 m3 /day(24-h residential inhalation rate) × 7 days

(16.1)

USEPA used this adjusted LOAEL of 2.5 mg/m3 to develop a reference concentration (RfC) of 3 µg/m3 [34]. The RfC is an exposure concentration associated with minimal, if any, risk for adverse effects, even in susceptible populations. The RfC was developed from the LOAEL by applying uncertainty factors to address use of a LOAEL instead of a no observed adverse effect level (NOAEL) (uncertainty factor of 10); use of a subchronic, rather than a chronic study (uncertainty factor of 3); intraspecies variability (uncertainty factor of 10); and database deficiencies, specifically, lack of chronic and multigenerational reproduction studies (uncertainty factor of 3). It should be noted that this RfC may be particularly conservative. For example, because cyanide does not accumulate in the body, the uncertainty factor for extrapolating from subchronic to chronic exposure may be unnecessary.

16.3.1.2 Subchronic/Chronic Toxicity Criterion — Ingestion This section presents toxicity criteria for evaluating subchronic and chronic ingestion exposure to free cyanide. Several authors have presented subchronic and chronic toxicity criteria for ingestion of free cyanide: Shifrin et al., 1996 [14]. These authors identified a NOAEL of 8 mg CN/kg-day, based on no changes in liver or body weight, in rats fed potassium cyanide (KCN) in the diet (as 200 µg KCN/g feed) for 21 days [35]. By applying a total uncertainty factor of 100 (10 for interspecies variability and 10 for intraspecies variability), Shifrin et al. derived a subchronic reference dose (RfD) of 0.08 mg/kg-day. USEPA, 1993 [36]. USEPA identified a NOAEL of 10.8 mg CN/kg-day, based on lack of effects on growth, no observable toxicity (e.g., hematological effects), and no histopathological changes, in a study in rats fed food fumigated with HCN (resulting in concentrations in food of 100 and 300 ppm) for 2 years [37]. By applying a total uncertainty factor of 100 (10 for interspecies variability and 10 for intraspecies variability) and a modifying factor of 5 (to account for the increased tolerance to cyanide when ingested with food), USEPA derived a chronic RfD of 0.02 mg/kg-day. ATSDR, 1997 [22]. ATSDR identified a NOAEL of 4.5 mg CN/kg-day, based on lack of reproductive effects in a study of male and female rats that ingested free cyanide in drinking water for 13 weeks [27]. In this study, rats were exposed to sodium cyanide in water at concentrations of 0, 3, 10, 30, 100, or 300 ppm (resulting in doses of 0, 0.2, 0.5, 1.7, 4.9, or 12.5 mgCN/kg-day in females, and 0, 0.2, 0.5, 1.4, 4.5, or 12.5 mg CN/kg-day in males) and evaluated for histopathology, clinical chemistry, hematology, urine chemistry, and reproductive toxicity. By applying a total uncertainty factor of 100 (10 for interspecies variability and 10 for intraspecies variability), ATSDR developed a minimal risk level (MRL), a toxicity criterion very similar to the RfD, of 0.05 mg/kg-day.

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It is interesting to note that the NOAELs from all three studies are within a factor of 2, despite the range in study durations (21 days to 2 years), endpoints (reproductive effects, observable toxicity, histopathology), and exposure media (water vs. food). This observation is consistent with a lack of accumulation of cyanide or its effects over time. Consistent with the potential for greater toxicity of cyanide in water than in food, the lowest NOAEL is from a drinking water study, although the difference is modest, possibly because drinking water is frequently ingested with food. For purposes of this analysis, we used the toxicity criterion of 0.05 mg CN/kg-day developed by ATSDR [22] because of the quality of the underlying study, the range in doses selected, the range of parameters measured by the investigators, and the use of drinking water as an exposure medium.

16.3.1.3 Acute Toxicity Criterion — Ingestion A pica child may potentially experience a single event exposure to a relatively high concentration of free cyanide in soil (i.e., a “hot spot”). Thus, it is relevant to consider use of a toxicity criterion for an acute ingestion exposure. Although the characteristics of acute cyanide poisoning are well described, there are limited data regarding single dose levels of cyanide that produced no effects in humans (i.e., a NOAEL). In a review of the literature on acute poisonings, the Massachusetts Department of Environmental Protection (MADEP) [38] identified 0.5 mg/kg as the lowest reported lethal dose of free cyanide, from a study of suicide victims by Gettler and Baine [39]. MADEP then extrapolated from this dose of 0.5 mg/kg to an allowable one time absorbed dose of 0.01 mg/kg, through application of a total uncertainty factor of 50 (5 for the use of a LOAEL, rather than a NOAEL, and 10 for intraspecies variability). Oddly, the LOAEL of 0.5 mg/kg for lethality in humans from the single dose analysis by Gettler and Baine [39] is less than the NOAEL for mild effects in animals from subchronic dosing in food or water, as observed in multiple studies. This discrepancy may reflect differences in response to cyanide ingested in a single dose vs. cyanide ingested throughout the day, as well as the ability of food to slow down the absorption, and hence toxicity, of free cyanide. However, this discrepancy also casts doubt on the reliability of the estimates of the lethal doses. For example, Gettler and Baine [39] estimated an absorbed dose of 24 mg HCN for the suicide victim in question, by extrapolating from the amount of HCN found in the liver and the brain, and subsequently estimated the amount ingested as 30 mg HCN, based on the amount of HCN remaining in the GI tract. Hence, the exact amount of cyanide ingested by the suicide victim was not known. In addition, the concentration of cyanide in the brain of the suicide victim in question was approximately 4- to 8-fold lower than brain concentrations of cyanide in other suicide victims, as well as dogs, that received lethal doses of cyanide [39]. This suggests that other factors may have contributed to the death of this suicide victim, possibly resulting in a lower threshold for poisoning by cyanide. Nonetheless, the estimated value of 30 mg ingested by the suicide victim agrees reasonably well with minimum lethal doses reported by other investigators of 50 to 60 mg [39], as well as a more recently reported value of 50 mg, or approximately 0.7 mg/kg, for a 70 kg individual [40].

16.3.2 FERRIC FERROCYANIDE SOLID Ferric ferrocyanide solid is noteworthy for its low toxicity. For example, no LD50 has been reported for ferric ferrocyanide. Unpublished data discussed by Giese [41], indicate that the LD50 for ammonium ferric ferrocyanide is well above 5000 mg/kg, a value greater than the LD50 of iron itself [41]. In fact, there was no evidence of hematological or histopathological effects in rats exposed via the diet to a dose of 4800 mg/kg-day for a period of four weeks [42]. Virtually all relevant studies that we have been able to identify on ferric ferrocyanide showed no evidence of toxicity [41,43]. In humans, constipation is the primary adverse reaction observed in adults receiving doses of 1.5 to 10 g/day (approximately 20 to 140 mg/kg-day for a 70 kg individual), for periods of 22 to 30 days [44,45].

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The low toxicity of ferric ferrocyanide solid is likely due both to its limited absorption from the GI tract, which is less than 0.1%, and its limited dissociation to free cyanide, either in the GI tract or in the bloodstream [46]. In other words, the amount of free cyanide in the bloodstream after ingestion of ferric ferrocyanide solid (i.e., bioavailability) is low. Nielsen et al. estimated that approximately 0.22% of cyanide ingested as ferric ferrocyanide solid will be available as free cyanide in the bloodstream. Bioavailability of the more soluble potassium ferric ferrocyanide (KFe[Fe(CN)6 ](s)) is only slightly greater than that of ferric ferrocyanide solid, with an estimated 0.26% of cyanide ingested as KFe[Fe(CN)6 ](s) available as free cyanide in the bloodstream [46]. Bioavailability of free cyanide in the bloodstream following ingestion of KFe[Fe(CN)6 ](s) is similar in humans, with approximately 0.27% of cyanide ingested as KFe[Fe(CN)6 ](s) available as free cyanide in the bloodstream [47]. Because of its limited systemic absorption and its ability to effectively bind certain cations, ferric ferrocyanide has been used therapeutically in Europe at doses of about 250 mg/kg-day, as an antidote for thallium poisoning [48]. The U.S. Food and Drug Administration (USFDA) has specified the conditions under which ferric ferrocyanide can be used for treatment of internal contamination with radioactive cesium, radioactive thallium, or nonradioactive thallium [49]. Doses up to 3 g/day are recommended for children of age 2 to 12 years, for a minimum treatment period of 30 days [50]. For a 15 kg child, this would be equivalent to 200 mg/kg-day.

16.3.2.1 Subchronic/Chronic Toxicity Criterion — Ingestion Dvorak and coworkers administered colloidal ferric ferrocyanide suspended in drinking water at 20,000 mg/l for 12 weeks to rats (equivalent to 3200 mg/kg-day), and observed no signs of toxicity, as assessed by histopathological examination of the GI tract, kidney, spleen, and liver [43]. By applying a total uncertainty factor of 100 to this NOAEL (10 for interspecies variability and 10 for intraspecies variability), we estimated a subchronic RfD for ferric ferrocyanide of 32 mg/kg-day. This value was also used as the chronic RfD, because ferric ferrocyanide does not accumulate over time.

16.3.2.2 Acute Toxicity Criterion — Ingestion There are no data available for developing an acute toxicity criterion for ferric ferrocyanide. However, for perspective, 3 g/day (equivalent to 200 mg/kg-day for a 15 kg child), is the dose recommended by the USFDA as an antidote for thallium poisoning, as discussed above [49].

16.3.3 DISSOLVED FERROCYANIDE As with ferric ferrocyanide solid, dissolved ferrocyanide also appears to have low toxicity. For example, Laforge et al. [51] discuss a patient who had reported ingesting a large amount of cyanide, which was confirmed by detecting toxic levels of cyanide in the blood, without manifesting any clinical signs of cyanide toxicity. The discrepancy between the cyanide exposure and the clinical manifestation was attributed to the form of cyanide ingested, which was later determined to be dissolved ferrocyanide. The low toxicity of soluble ferrocyanide is likely due to its low absorption, limited cellular uptake, rapid elimination from the body, and limited ability to generate free cyanide in the GI tract or the bloodstream [46,52]. Nielsen et al. [46] estimated that less than 3% of ferrocyanide in the GI tract is absorbed into the bloodstream. Although soluble ferrocyanide was found to bind to plasma proteins, it did not appear to partition into red blood cells, and was not detected in gastric juice, feces, or saliva [52]. In four normal human subjects (without kidney disease), 68 to 87% of an intravenous dose was eliminated via the urine within 24 to 48 h [52].

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16.3.3.1 Subchronic/Chronic Toxicity Criterion — Ingestion There are no suitable studies available for deriving a toxicity criterion for soluble ferrocyanide, that is, studies with well-defined effects other than mortality, and a dose–response relationship. However, the study by Dvorak and coworkers [43], discussed above, can be used to estimate a toxicity criterion for dissolved ferrocyanide. Specifically, the estimated concentration of dissolved ferrocyanide in the Dvorak study, based on estimated dissolution by the investigators, would be 1400 mg/l, which is equivalent to an ingested dose of approximately 200 mg/kg-day. By applying a total uncertainty factor of 100 as above (10 for interspecies variability and 10 for intraspecies variability), the estimated subchronic and chronic RfD for dissolved ferrocyanide would be 2 mg/kg-day.

16.3.3.2 Acute Toxicity Criterion — Ingestion In contrast to the soil pathway, where the possibility of an acute exposure due to pica soil ingestion exists, acute exposure to large amounts of cyanide in water is much less likely to occur (if at all). The reason for the low probability of water exposure resulting in an acute risk of dissolved ferrocyanide is that there is no drinking water equivalent of pica soil ingestion. That is to say, water intake cannot exceed certain physiological limits without adversely affecting the individual. The following theoretical example demonstrates how acute exposures are not relevant to water exposure of dissolved ferrocyanide. The toxicity benchmark for this comparison, 1600 mg/kg, is the lowest oral dose producing mortality in rats [53]. Given our assumed concentration of 11 mg/l of dissolved ferrocyanide in water, a 70 kg individual would have to ingest the highly unlikely amount of more than 10,000 liters of water to approach this dose, as follows: 1600 mg/kg × 70 kg = 10, 812 liters 11 mg/l

(16.2)

Given this unlikely acute scenario, we did not evaluate quantitatively acute exposure to dissolved ferrocyanide in this risk analysis.

16.4 RISK CHARACTERIZATION This section presents screening level risk estimates, using upper bound parameters for concentrations in environmental media and for intake estimates. While the selected parameters are individually plausible, it is unlikely that all such parameters would exist together at a single site. For example, we have assumed a subchronic exposure scenario in which a young child (age 1 to 6 years) is exposed to the assumed soil, water, and air concentrations on a daily basis. This is an unlikely scenario, considering that many MGP sites are on industrial rather than residential properties, and that exposures to nearby residents would tend to involve trespasser situations of adolescents, with less frequent and less intense exposure. In addition, it should be recognized that all risk assessments contain elements of variability (i.e., ranges in selected parameters or outputs) and uncertainty (i.e., lack of true knowledge regarding a specific parameter). Variability and uncertainty may be characterized through a range of approaches, including qualitative descriptions of uncertainty, sensitivity analyses with use of alternate parameters, and probabilistic modeling [54]. Because our aim here is to provide a general methodology for plausible “worst case” illustrative purposes, no formal uncertainty analysis was conducted, although we will identify key sources of uncertainty and their potential significance. As with any risk assessment, there are several sources of uncertainty in this analysis. Overall, where there is uncertainty, values were selected that would likely overestimate risk. Key sources of

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uncertainty in this analysis include estimation of air concentrations, evaluation of dermal uptake, and selection of an acute toxicity criterion for free cyanide. In the absence of sufficient data for estimating typical air concentrations of free cyanide at MGP sites, air concentrations were modeled using several conservative assumptions regarding emission rates. For example, we assumed a minimum half-life for cyanide compounds in soil, and that all cyanide compounds would be released in the gaseous form as HCN, both of which are assumptions that would overestimate emission rates. We also assumed a conservatively high concentration of 5000 mg/kg for ferric ferrocyanide solid in soil. Furthermore, we assumed that an individual would be exposed to the maximum modeled air concentration for 24 h/day, 350 days/year, whereas ambient air concentrations would not always be as high as the maximum (depending on directions of the prevailing winds). In addition, the individual would not spend all of his or her time outside, at the same location. Another key source of uncertainty relates to the evaluation of dermal uptake. In the absence of information regarding dermal absorption, we assumed that dermal absorption would be equivalent to oral absorption. However, as discussed earlier, dermal absorption would likely be much lower than oral absorption. Finally, there is also uncertainty regarding the selection of an acute toxicity criterion for free cyanide. This toxicity criterion was based on an older study of suicide victims who had ingested cyanide. A major uncertainty for this study is lack of information regarding the specifics of the suicide, such as details on whether the victim may have been exposed to other agents that could have contributed to the suicide.

16.4.1 EXPOSURE ALGORITHMS AND ASSUMPTIONS This section presents the algorithms used to calculate intake for the screening assessment, as well as for the specific intake parameters.

16.4.1.1 Soil For calculating exposure to soil via ingestion and dermal uptake, the following equations were used: Ingested dose (CN− or ferric ferrocyanide solid)in mgcompound /kgbody weight -day = (Cs × IRs × 10−6 kg/mg)/BW

(16.3)

where Cs is the concentration of compound in soil (mgcompound /kgsoil ), IRs the soil ingestion rate (mgsoil /day), and BW the body weight of the child (kg). Dermal dose (CN− or ferric ferrocyanide solid) in mgcompound /kgbody weight -day = (Cs × SAs × AF × 10−6 kg/mg)/BW

(16.4)

where Cs is the concentration of compound in soil (mgcompound /kgsoil ), SAs the skin surface area exposed to soil (cm2 ), AF the skin–soil adherence factor (mgsoil /cm2skin ), and BW the body weight of the child (kg). The soil concentrations, as well as the values and basis for the specific intake parameters, are presented in Table 16.4.

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TABLE 16.4 Parameter Assumptions for Ingestion and Dermal Exposure to Soil at a Hypothetical MGP Site Assumption

Value

Concentration of free cyanide

5 mg/kg

Concentration of ferric ferrocyanide

5000 mg/kg

Concentration of free cyanide at hot spot

20 mg/kg

Soil ingestion rate — subchronic/chronic

100 mg/d

Soil ingestion — acute Skin surface area

1 g/event 2800 cm2

Skin–soil adherence factor

0.2 mg/cm2

Body weight

15 kg

Basis Assumed, based on data indicating that upper end of fraction of cyanide in iron complexes that is present as CN− is 0.1 % Assumed, based on upper end of range of 12 sites in Netherlands [8] Assumed, based on exposure to OBW at 20,000 mg/kg total cyanide, of which 0.1% is free cyanide Recommended daily soil ingestion rate [25] Professional judgment Skin surface area for face, hands, forearms, lower legs, and feet, for 1–6-year-old child [61] Weighted soil adherence factor assuming exposure to face, forearms, hands, lower legs, and feet [61] Mean body weight for 1–6-year-old child [61]

16.4.1.2 Groundwater For calculating exposure to groundwater via the ingestion and dermal route, the following equations were used: Ingested dose (CN− or dissolved ferrocyanide) in mgcompound /kgbody weight -day = (Cgw × IRgw )/BW

(16.5)

where Cgw is the concentration of compound in water (mgcompound /lwater ), IRgw the drinking water ingestion rate (lwater /day), and BW the body weight of the child (kg). Dermal dose (CN− or dissolved ferrocyanide) in mgcompound /kgbody weight -day = (Cgw × SAgw × ET × Kp × 10−3 l/cm3 )/BW

(16.6)

where SAgw is the skin surface area for bathing (cm2 ), ET the exposure time (h), and Kp the dermal permeability coefficient (cm/h). Water concentrations, as well as the values and basis for the specific intake parameters, are presented in Table 16.5. 16.4.1.3 Air For the air pathway, no intake was estimated. Rather, the estimated air concentration was compared to the RfC (described in Section 16.3.1.1). The predicted air concentration for HCN is 0.5 µg/m3 , as discussed earlier.

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TABLE 16.5 Parameter Assumptions for Ingestion and Dermal Exposure to Groundwater at a Hypothetical MGP Site Assumption

Value

Basis

Concentration of free cyanide

0.1 mg/l

Concentration of dissolved ferrocyanide

11 mg/l

Drinking water ingestion rate

1.1 l/day

Skin surface area for bathing

6600 cm2

Exposure time Dermal permeability of free cyanide and ferric ferrocyanide Body weight

1h 1 × 10−3

Assumed, based on upper end of range of 12 sites in Netherlands [8] Assumed, based on upper end of range of 12 sites in Netherlands [8] Recommended ingestion rate for drinking water [25] Total body skin surface area for 1–6 year-old-child [61] Recommended exposure time for bathing [61] Default for inorganics [61]

15 kg

Mean body weight for 1–6-year-old child [61]

16.4.2 CASE STUDY RESULTS Risk results for subchronic exposure in this case study are presented as hazard quotients (HQs). The HQ represents the ratio of the estimated exposure dose (described in Section 16.4.1) to the appropriate toxicity criterion (described in Section 16.1.3). A HQ below 1 indicates negligible concern for toxicity, even for susceptible populations. A HQ greater than 1 does not mean that adverse effects will necessarily occur (or even are likely), but rather, indicates that additional evaluation is needed. The HQs for this case study are presented in Table 16.6. Even using conservative assumptions for environmental media concentrations and for intake parameters, all HQs are below 1. The HQs range from 6.7 × 10−4 for ingestion of free cyanide in soil to 1.7 × 10−1 for inhalation of free cyanide in air. Summing the individual HQs across exposure pathways yields a total hazard index (HI) of 0.35 for free cyanide, less than 0.01 for ferric ferrocyanide solid, and 0.40 for dissolved ferrocyanide. Hence, the total HI (across pathways and cyanide species) is below the target of an acceptable HI of 1. Thus, under most conditions, risks from free cyanide and dissolved-phase and solid-phase iron cyanides at MGP sites are not expected to be significant. An analysis was also conducted for acute ingestion exposure to free cyanide in 1 g of soil from a hot spot of 20 mg free cyanide/kg soil, using the MADEP acute toxicity criterion for available cyanide of 0.01 mg/kg (where available cyanide is defined by MADEP as “those species of cyanide which are capable of releasing [HCN] or the cyanide anion [CN− ] under reasonably anticipated human gastric conditions,” [38]). The estimated intake from soil of 0.0013 mg/kg-day is about 7-fold below the toxicity criterion. Even if one were to assume a soil ingestion rate of 5 g/day (as recommended by ATSDR) [55], the estimated intake from soil would not exceed the toxicity criterion. As noted earlier, the acute toxicity criterion for free cyanide may be overly conservative, and thus, the acute HQ could be even lower. The risks that most closely approach the HQ of 1 are due to HCN in air, and HCN and dissolved ferrocyanide in groundwater. Had the cyanide concentrations in water been erroneously assumed to consist entirely of free cyanide, all HQs would have been well above 1, and would have indicated a toxicological concern. Thus, this screening analysis confirms the need to estimate free cyanide concentrations, preferably through measurements, or, if measurements are not available, through the use of realistic assumptions regarding the likely fraction of free cyanide present in soil and groundwater.

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TABLE 16.6 Potential Risks Associated with Subchronic and Chronic Ingestion and Dermal Exposures to Cyanide Compounds at a Hypothetical MGP Site Pathway Ingestion of soil Dermal contact with soil Ingestion of groundwater Dermal contact with groundwater Inhalation of air Total hazard index

Free cyanide

Ferric ferrocyanide solid

Dissolved ferrocyanide

0.00066 0.0037 0.15 0.022 0.17 0.35

0.001 0.0058 Not applicable Not applicable Not applicable 0.0068

Not applicable Not applicable 0.4 0.001 Not applicable 0.401

Total hazard index for free cyanide, ferric ferrocyanide solid, and dissolved ferrocyanide

0.76

TABLE 16.7 Selected Soil Criteria for Ingestion of Cyanide in Soil Agency

Criterion (mg/kg)

Form of CN in soil

Basis

Reference

ILEPA MADEP NJDEP USEPA

1600 100 1100 1600

Amenable “Available”a Total Amenable

Chronic Acute Chronic Chronic

[62] [38] [63] [64]

a Available cyanide defined as “forms of cyanide capable of releasing HCN or CN− in the gut.”

16.4.3 REGULATORY CRITERIA FOR CYANIDE COMPOUNDS IN ENVIRONMENTAL MEDIA 16.4.3.1 Soil Table 16.7 lists some risk-based regulatory criteria for cyanide in soil for ingestion exposure. This table is not meant to be comprehensive, but rather, is meant to illustrate certain issues relating to assessment of risks from cyanide exposure. The four criteria presented recommend that cyanide should be characterized in three different ways: amenable cyanide (USEPA, ILEPA), physiologically available cyanide (MADEP), or total cyanide (NJDEP). Amenable cyanide is cyanide amenable to chlorination, and represents total cyanide minus iron-complexed species [14]. (The long-used amenable cyanide analytical technique has a number of systemic problems, and is not the best available measure of free cyanide and weak metal–cyanide complexes, as discussed in Chapter 7.) Physiologically available cyanide (referred to by MADEP as “available cyanide,” described in Section 16.4.2) is defined by MADEP as “. . . those species of cyanide which are capable of releasing Hydrogen Cyanide (HCN) or the cyanide anion (CN− ) under reasonably anticipated human gastric conditions” [38]. (Note that the MADEP definition of “available cyanide” differs from available

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cyanide as determined by the USEPA-approved Available Cyanide Method [OIA-1677], which measures free cyanide and weak metal–cyanide complexes, such as cyanide complexes with Cu, Zn, Ni, Cd, Ag, and Hg.) Three of the criteria (USEPA, ILEPA, NJDEP) are based on chronic exposure, whereas one (MADEP) is based on a single acute exposure. The most restrictive criterion (100 mg/kg) is based on acute exposure, and is more than an order of magnitude lower than criteria based on chronic exposure, which are all similar in magnitude (1100 to 1600 mg/kg). It should be noted that the acute soil criterion of 100 mg/kg was based on the acute toxicity study discussed in Section 16.3.1.3, and thus may not be as reliable as the other values. This comparison of cyanide soil criteria illustrates the need for evaluation of the appropriate species of cyanide in soil. Using the case study of the MGP site presented earlier in this chapter (which assumed a total cyanide concentration of 20,000 mg/kg and a free cyanide concentration of 5 mg/kg), cyanide concentrations would have exceeded the NJDEP criterion of 1100 mg/kg (based on total cyanide), but not the USEPA or the ILEPA limit of 1600 mg/kg (based on amenable cyanide). It is difficult to compare the case study results with the MADEP criterion, which is based on a characterization of cyanide that corresponds to an analytical methodology still under development. For MGP sites, where most of the cyanide present is in the form of ferric ferrocyanide solid [14], a health criterion based on total cyanide is clearly inappropriate. A measure of free cyanide would be more appropriate on the soil leachate, or, by the MADEP definition, cyanide that either is free in the environment or could be converted to free cyanide in the GI tract. From the risk characterization standpoint, the choice of the appropriate analytical method to quantify the toxicological species of concern in the exposure media, that is, soil and water, is very important. In that regard, the free cyanide method by microdiffusion [56] is a good choice, as far as detection of free cyanide in water or soil leachates is concerned (see Chapters 7 and 8). Other conventional analytical techniques, such as cyanide amenable to chlorination or weak-acid-dissociable cyanide (both are described in Standard Methods as Method 4500-CN [57], with the latter preferred as being less subject to interferences), tend to overestimate the amount of free cyanide exposure from water because of two reasons: (1) these methods can recover weak metal–cyanide complexes involving Cd, Cu, Ni, Ag, and Zn, in addition to biologically available free cyanide (HCN or CN− ) in water matrices, and (2) not all of the noted metal–cyanide complexes will dissociate into free cyanide in the GI tract, once the sample is ingested. For more details on cyanide analytical methods in solid and water media, the reader should refer to Chapters 7 and 8, respectively. It is also of interest to note that the soil criterion for an acute, single exposure is more restrictive than the soil criterion for subchronic or chronic exposure. This is a function of the more restrictive toxicity criterion for a single event exposure (0.01 mg/kg-day, as compared to the chronic toxicity criterion of 0.02 mg/kg-day) as well as the higher soil ingestion rate used for acute exposures. While MADEP notes that it may be counter-intuitive for an acute RfD (and, by inference, associated soil cleanup level) to be more restrictive than a chronic RfD, it concludes that such an anomaly could be explained in the case of cyanide by the body’s ability to detoxify cyanide when it is ingested in multiple doses, vs. a single bolus dose. However, this anomaly may also be due to uncertainties in the acute toxicity criterion itself (discussed in Section 16.3.1.3).

16.4.3.2 Water The USEPA maximum contaminant level (MCL) for cyanide in drinking water is 0.2 mg/l. This limit is based on free cyanide [58]. The value was developed in 1989, based on an equivalent NOAEL in drinking water of 0.76 mg/l from a study by Howard and Hanzal [37], and assuming a relative source contribution from drinking water of 20% [59]. However, it is not uncommon for cyanide in water to be measured as total cyanide (e.g., as is done by the Indiana Department of Environmental Management) [60].

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In our case study, analysis of total cyanide in groundwater would have resulted in a significant exceedance of the MCL, suggesting a potentially significant health risk. However, the analysis of more appropriate cyanide species, as discussed above, would have demonstrated that the site did not present a significant health risk in terms of ingestion of, and dermal exposure to, groundwater. A perspective on measurements of cyanide in water may be provided by comparison of cyanide levels in certain foods. Cyanide may be present as HCN in juices from stone fruits (e.g., peaches, apricots, prunes, cherries), at concentrations ranging from 1.9 to 4.6 mg/l [22]. While individuals typically consume less fruit juice than water, by volume, it is striking, nonetheless, that on average, the concentration in juices that are consumed by individuals in the U.S. population is more than 10-fold greater than the MCL.

16.5 RECOMMENDATIONS This section presents recommendations for conducting risk assessments for exposure to cyanide at MGP sites. As discussed above, free cyanide, which is substantially more toxic than the iron-complexed forms of cyanide typically found at MGP sites, represents only a small fraction of the total cyanide at these sites, with most of the cyanide in the form of ferric ferrocyanide solid. Therefore, risk assessments should be based on cyanide species concentrations, with free cyanide of primary interest, rather than total cyanide. Information on weak-acid-dissociable cyanide, available cyanide, strongly-complexed cyanide, and individual metal–cyanide species will also be helpful. In addition, while there is information available regarding absorption of both free cyanide and ironcomplexed cyanide from the GI tract, there is little information regarding absorption of free cyanide or ferric ferrocyanide when present in soil (i.e., soil bioavailability). As such, we recommend that studies should be conducted to determine bioavailability of free cyanide, metal–cyanide species, and ferric ferrocyanide in soil. Our analysis used modeled concentrations of free cyanide in air, based on concentrations of free cyanide in soil. However, because of uncertainties in a range of factors, such as the release rate of free cyanide from soil, quantification of free cyanide concentrations in air would provide more precise estimates of inhalation exposures. Therefore, we recommend, where feasible, air monitoring for free cyanide concentrations at MGP sites.

16.6 SUMMARY AND CONCLUSIONS • Cyanide compounds are widely used in industry, and their release can pose human health risks. The occurrence of cyanide in water, soil, and air at the thousands of former MGP sites in the United States provides a good case study for assessment of the human health risk associated with environmental release of cyanide. • Cyanide at MGP sites has low toxicity overall, with much of the total cyanide present in relatively nontoxic forms of ferrocyanide, such as ferric ferrocyanide solid and dissolved ferrocyanide, and only a small fraction present in the more toxic form of free cyanide. • Free cyanide can be highly toxic to the brain and central nervous system, with a minimum lethal dose of 0.7 mg/kg. In contrast, humans can be exposed to ferric ferrocyanide solid at ingestion doses as high as 20 to 140 mg/kg for at least 22 days, with minimal effects. For therapeutic purposes, ferric ferrocyanide can be ingested at doses up to 200 mg/kg-day for 30 days. • The low toxicity of ferric ferrocyanide relates to its low volatility and low absorption across biological membranes, with less than 0.1% absorbed through the GI tract. Less than 0.3% of cyanide ingested as ferric ferrocyanide is available as free cyanide in the bloodstream.

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• Based on a conservative screening analysis, exposures to cyanide in soil and groundwater at MGP sites present a negligible toxicological concern, even for susceptible subpopulations. This analysis was conducted using very high concentrations of total cyanide and involved multiple exposure pathways, including ingestion of, dermal contact with, and inhalation of cyanide from both soil and groundwater. • The results from this screening analysis demonstrate the importance of distinguishing between free and strongly-complexed cyanide at MGP sites and in other cyanide release cases, due to the considerable difference in toxicity between these different forms of cyanide. • In order to reflect more accurately the risks from exposure to cyanide compounds typically found at MGP sites, cyanide concentrations should be quantified as both free cyanide and iron-complexed cyanide. • Bioavailability studies of available free cyanide, metal–cyanide species, and ferric ferrocyanide in soil should be conducted in order to define better the systemic absorption of free cyanide from ingested soil. • Where feasible, measurement of air concentrations of free cyanide should be considered, in order to characterize more accurately the risk from inhalation exposures.

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38. MADEP, Background documentation for the development of an “available cyanide” benchamrk concentration, Massachusetts Department of Environmental Protection, http://www.mass.gov/dep/ord/files/ cn_soil.htm,1998, accessed: July 27, 2004. 39. Gettler, A.O. and Baine, J.O., Toxicology of cyanide, Am. J. Med. Sci., 195, 182, 1938. 40. Ellenhorn, M.J., Ellenhorn’s Medical Toxicology: Diagnosis and Treatment of Human Poisoning, 2nd ed., Williams & Wilkins, Baltimore, MD, 1997. 41. Giese, W.W., Ammonium-ferric-cyano-ferrate(II) (AFCF) as an effective antidote against radiocaesium burdens in domestic animals and animal derived foods, Br. Vet. J., 144, 363, 1988. 42. Kostial, K., Kargacin, B., Rabar, I., Blanusa, M., Maljkovic, T., Matkovic, V., and Ciganovic, M., Simultaneous reduction of radioactive strontium, cesium and iodine retention by single treatment in rats, Sci. Total Environ., 22, 1, 1981. 43. Dvorak, P., Gunther, M., Zorn, U., and Catsch, A., Metabolisches verhalten von kolloidalem ferrihexacyanoferrat(II)=Metabolic behavior of colloidal ferrihexacyanoferrate(II), Nauyn-Schmiedebergs Arch. Pharmak., 269, 48, 1971. 44. Brandao-Mello, C.E., Oliveira, A.R., Valverde, N.J., Farina, R., and Cordeiro, J.M., Clinical and hematological aspects of 137Cs: the Goiania radiation accident, Health Phys., 60, 31, 1991. 45. Madshus, K. and Stromme, A., Increased excretion of 137Cs in humans by Prussian Blue, Z. Naturforsch B., 23, 391, 1968. 46. Nielsen, P., Dresow, B., Fischer, R., and Heinrich, H.C., Bioavailability of iron and cyanide from 59 Fe- and 14 C-labelled hexacyanoferrates(AA)in rats, Z. Naturforsch B., 45C, 681, 1990. 47. Nielsen, P., Dresow, B., Fischer, R., and Heinrich, H.C., Bioavailability of iron and cyanide from oral potassium ferric hexacyanoferrate(II) in humans, Arch. Toxicol., 64, 420, 1990. 48. Spoerke, D.G., Smolinske, S.C., Wruk, K.M., and Rumack, B.H., Infrequently used antidotes: indications and availability, Vet. Hum. Toxicol., 28, 69, 1986. 49. FDA, Guidance for industry on Prussian Blue for treatment of internal contamination with thallium or radioactive cesium; availability, Docket No. 03D-0023, U.S. Food and Drug Administration, Washington, DC, Federal Register, 68(23), 2003. 50. USFDA, Draft labeling for insoluble Prussian Blue capsules, Center for Drug Evaluation and Research, http://www.fda.gov/cder/foi/label/2003/ind51700lbl.pdf, accessed: July 16, 2004. 51. Laforge, M., Gourlain, H., Fompeydie, D., Buneaux, F., Borron, S.W., and Galliot-Guilley, M., Ferrocyanide ingestion may cause false positives in cyanide determination, Clin. Toxicol., 37, 337, 1999. 52. Kleeman, C.R. and Epstein, F.H., Fate and distribution of Fe59 labelled ferrocyanide in humans and dogs, Proc. Sco. Exp. Biol. Med., 93, 228, 1956. 53. NIOSH, Registry of Toxic Effects of Chemical Substances (RTECS) record for sulfuric acid, CAS #7664-93-9, National Institute of Occupational Safety and Health, 2000. 54. Ramaswami, A., Milford, J.B., and Small, M.J., Integrated Environmental Modeling: Pollutant Transport, Fate and Risk in the Environment, John Wiley & Sons, New York, 2005. 55. ATSDR, Summary report for the ATSDR soil-pica workshop June 2000, contract no. 205-95-0901, Agency for Toxic Substances and Disease Registry, Atlanta, GA, http://www.atsdr.cdc.gov/NEWS/ soilpica.html, 2001, accessed: October 17, 2003. 56. ASTM, Standard test method for determination of free cyanide in water and wastewater by microdiffusion, Method D-4282-02, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2004. 57. APHA/AWWA/WEF, Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S., Greenberg, A.E., and Eaton, A.D., Eds., American Public Health Association, American Water Works Association, and Water Environment Federation, Washington, DC, 1998. 58. USEPA, National primary drinking water standards and national secondary drinking water standards, EPA/816-F-03-016, U.S. Environmental Protection Agency, Office of Water, Washington, DC, http://www.epa.gov/safewater/consumer/mcl.pdf, 2003. 59. USEPA, Proposed rules 40 CFR Parts 141, 142 and 143 (WH-FRL-3701-9) RIN 2040-AB11: National primary and secondary drinking water regulations; synthetic organic chemicals and inorganic chemicals, Fed. Regist., 55, 30370, 1990.

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Risk Assessment of 17 Ecological Cyanide in Water and Soil Roman P. Lanno and Charles A. Menzie CONTENTS 17.1

Problem Formulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.1.1 Conceptual Models of Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.1.1.1 Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.1.1.2 Receptors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.1.2 Assessment Endpoints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.1.2.1 Assessment Endpoints for Aquatic Environments . . . . . . . . . . . . . . . . . . 17.1.2.2 Assessment Endpoints for Terrestrial Environments . . . . . . . . . . . . . . . . 17.2 Ecological Effects Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2.1 Metabolic Basis of Cyanide Reactivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2.2 Lethality of SCN− and CN− . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2.3 Sublethal Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2.4 Comparative Sublethal Toxicity of CN− and SCN− . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2.5 Species Sensitivity Distributions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3 Ecological Exposure Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3.1 Aquatic Environments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3.2 Terrestrial Environments. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4 Risk Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

332 332 335 336 336 336 337 337 337 338 339 339 340 343 343 344 345 346 346

Cyanide in water and soil may pose significant risk for ecological receptors, depending upon the chemical species of cyanide and the nature of exposure. This chapter outlines the key aspects of exposure and effects that influence the ecological risk associated with releases of cyanide. The chapter is organized in a format that follows the standard framework for risk assessment [1,2]: problem formulation, exposure assessment, effects assessment, and risk characterization. Due to the many data gaps that exist in the literature, this summary of cyanide-related risk does not provide a complete picture. However, it does provide an overview of the current state of the science as it applies to the completion of an ecological risk assessment for cyanide. Existing data gaps and research needs that may reduce uncertainty are discussed. It is helpful to begin with some key points that shape the discussion of ecological risk. Eisler [3] notes that most authorities agree on five points with respect to exposure to mammals, which also apply in varying degrees to fish and other wildlife: 1. Toxic inorganic forms of cyanide, like free cyanide ion, CN− , and molecular HCN, have low persistence in the environment (surface soils, sediments, and surface water), and are 331

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2. 3. 4.

5.

not accumulated or stored in any mammal (or other organisms); toxic forms of inorganic cyanide can be persistent in certain wastes and in groundwater. Cyanide biomagnification in food webs has not been reported, possibly due to rapid detoxification of sublethal doses by many species, and death at higher doses. Cyanide has an unusually low occurrence of chronic toxicity (in mammals) but chronic intoxication exists, and in some cases, can be incapacitating. Despite the high lethality of large single doses or acute respiratory exposures to high vapor concentrations of cyanide (in wildlife), repeated sublethal doses seldom result in cumulative adverse effects. Intermittent high doses (to wildlife) can be tolerated by many species for long periods, without adverse effects.

Organocyanides can be produced and stored by some plants, animals, and microbes (see Chapter 3). Exposure to these naturally produced organic cyanide compounds can have lethal or sublethal effects. While we occasionally refer to these sources of cyanide, this chapter focuses on the inorganic cyanides that occur as waste products from industrial operations. Applications of cyanide-based pesticides have posed risks to wildlife receptors, especially bird species, that inhabit treated areas [3]. Similarly, the use of cyanide poisons to collect fish from coral reefs can also decimate local fish populations and destroy coral reefs ([4]; and Chapter 11). These sources of cyanide will be discussed briefly, since it is clear that they do have deleterious effects on ecological receptors.

17.1 PROBLEM FORMULATION Problem formulation involves defining the problem to be addressed, the development of assessment endpoints (what are we trying to protect?), and the development of a plan for analyzing and characterizing the risk. During this phase of risk assessment, hypotheses and conceptual models are developed, and assessment endpoints and measures of effects are identified.

17.1.1 CONCEPTUAL MODELS OF EXPOSURE A conceptual model is a narrative and visual depiction of predicted relationships between ecological entities and the stressors to which they may be exposed [5]. It guides the technical and managerial approaches for addressing that problem. Cyanide exposures can be associated with releases to air, soil, groundwater, and surface water. The environmental fate of cyanide can involve movement among these media (see Chapters 9 to 11). For example, releases from soil or solid waste to groundwater can also result in subsequent discharges to surface water bodies. Conceptual model diagrams help capture the relationships between sources, transport pathways, and human or ecological receptors in a way that helps frame the problem at hand. Conceptual models consist of two principal components [5]: (1) a set of risk hypotheses that describe predicted relationships among stressors, exposure, and assessment endpoint response, along with the rationale for their selection; (2) a diagram that illustrates the relationships among sources, pathways, and receptors. Conceptual model diagrams are used to illustrate important pathways clearly and concisely, and can be used to generate new questions about relationships that help formulate plausible risk hypotheses. Typical conceptual model diagrams are flow diagrams containing boxes and arrows to illustrate relationships. In this chapter, we focus on graphical conceptual models to illustrate the features of such models for addressing cyanide contamination problems. We do not illustrate or discuss specific site scenarios

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Anthropogenic cyanide sources

Natural cyanide sources

Anthropogenic discharges

Microbial cyanide sources

Cyanide in plant tissues

Inorganic cyanide

Water column

Leaching pits

Soil content Wildlife exposure

Fish and invertebrate exposure

Fish and invertebrate effects

Soil porewater

Soil air

Groundwater

Surface air

Wildlife effects

FIGURE 17.1 Conceptual model for cyanide ecological risk analysis.

and related risk hypotheses, although we do identify some expected relationships and potential effects that are reflected in the development of assessment endpoints. We begin with a general conceptual model for cyanide in the environment (Figure 17.1). This shows a variety of possible exposure pathways. As discussed in Chapters 9 to 12, some of these pathways will have high relevance for anthropogenic sources, while others will be less important. For these anthropogenic sources, we illustrate the major and less important pathways in aquatic (Figure 17.2) and terrestrial (Figure 17.3) ecosystems. We include both major and less important pathways because much of the uncertainty around exposure is related to an understanding of exposure pathways. Aquatic and terrestrial systems receive cyanide inputs from both anthropogenic and natural sources. Major anthropogenic activities associated with releases of elevated concentrations of free and metal-complexed cyanide, either through effluent discharges or through surface runoff and solid waste disposal, include gold mining, metal extraction, electroplating and metal finishing, petrochemical refineries, power plants, and the combustion of solid wastes, both industrial and domestic (see Chapters 4,25,27). Total cyanide concentrations in wastewaters and solid residuals from these processes can range from ppb levels to ppm levels in the tens of thousands. Solid waste residuals from some industries (e.g., spent potliner from aluminum smelting and spent oxide box waste from former manufactured gas plant [MGP] facilities) that were land-disposed can be sources of cyanide exposure for terrestrial ecological receptors. Some of these solid waste streams can contribute to surface water exposure, either via runoff or through the groundwater pathway (see Chapters 9 and 10). The major cyanide-bearing waste at MGP sites is in the form of ferric ferrocyanide (Fe4 [Fe(CN)6 ]3 (s)), with thiocyanates and soluble simple cyanides present at low concentrations. Liquid wastes produced from numerous industries (e.g., gold mining, coke plants, and petroleum refineries, etc.) may contain some forms of cyanide. Cyanide is also used in the extraction of precious metals from heaps of finely crushed ore (see Chapters 4 and 26). Leach heaps may be 1 ha in size, with

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Sources Air Point sources

Fate processes

Aquatic receptors

Photolysis Oxidation Advection Mixing

Fish

Wildlife receptors

Wildlife that eat fish

Surface water Stormwater & nonpoint sources

Wildlife that eat invertebrates or plants

Surface sediment (biologically active zone)

Spills Burial

Invertebrates -------Plants

Deep sediment Groundwater Sorption Desorption Degradation

Often important Sometimes important Less important

FIGURE 17.2 Conceptual model for cyanide ecological risk analysis in aquatic ecosystems. Note that all of the processes shown have specific spatial and temporal scales.

Sources

Deposits of solid and semi-solid wastes

Spills

Soil receptors

Fate processes Bioturbation Scouring Deposition Resuspension Transport

Plants

Surface soils biologically active zone Soil invertebrates

Stormwater & nonpoint sources

Wildlife receptors

Wildlife that eat plants

Wildlife that eat invertebrates

Fill & Cover

Inaccessible soils

Often important Sometimes important Less important

Sorption Desorption Degradation

FIGURE 17.3 Conceptual model for cyanide ecological risk analysis in terrestrial ecosystems. Note that all of the processes shown have specific spatial and temporal scales.

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puddles on the tops of heaps and exposed solution in recovery channels at the base of heaps representing the highest concentrations of cyanide. Wastewater from this process is often stored in ponds, which can represent potential significant hazards to wildlife [6,7]. Holding ponds and other open tankage are often used in management of other industrial wastewaters and contaminated groundwaters, and these can also represent potential exposure points for aquatic organisms and wildlife. Eisler et al. [7] document hazards to wildlife from wastewater management at gold mining operations, and suggest some mitigation options. A special case of aquatic release of cyanide involves a concentrated solution of NaCN for collection of specific, high value fish, especially in Southeast Asia (see Chapter 11). Similar methods have been used to collect fish for biological surveys. The use of cyanide administered as a concentrated solution from a squirt bottle for the capture of marine tropical fish by divers has spread throughout Southeast Asia, in association with the live fish trade for marine aquarium fish, and for groupers exported to Hong Kong and mainland China as food fish [4,8–13]. This has resulted not only in the acute mortality of fish, but also in the destruction of coral reefs. As described earlier and discussed further in Chapter 3, significant naturally-occurring levels of cyanide in terrestrial and aquatic systems may also arise within cyanogenic plants and animals, as the metabolic by-products of fungi, bacteria, and blue-green algae (Cyanophyta), and by the release and hydrolysis of cyanogenic glycosides during the decomposition of higher plants. These nonanthropogenic sources of cyanide contribute to existing background levels of cyanide in terrestrial and aquatic systems.

17.1.1.1 Pathways The fate and transport of cyanide in soils, groundwater, and surface waters are described in Chapters 9 to 11. Here, we consider characteristics of pathways that are important to exposure of aquatic and terrestrial ecological receptors (Figure 17.2 and Figure 17.3). For aquatic environments, exposure occurs through direct contact with water. As illustrated in Figure 17.2, exposures are greatest for animals and plants that are immersed in the water; wildlife species can be exposed by drinking water. Incidental ingestion of cyanides in sediment is not considered a significant pathway, because toxic forms of cyanide (free cyanide and weak metal–cyanide complexes) are not expected to accumulate there. Free cyanide exhibits little tendency to adsorb on sediments [14]. In sediment, free cyanide is either chemically converted to nontoxic cyanide-containing compounds (e.g., iron–cyanide solids), or metabolized to innocuous compounds (e.g., CO2 and ammonia), as described in Chapter 6. Importantly for assessing ecological risk, bioaccumulation and transfer of cyanide via the food web is not a significant pathway for potentially toxic inorganic cyanide compounds [3]. Some aquatic organisms can produce toxic cyanide compounds (i.e., they are cyanogenic), and the presence of organic cyanide compounds in the tissues of these organisms can be a source of exposure. However, these processes are not related to anthropogenic sources of cyanide, and possess no significant environmental impacts, except in unusual circumstances (e.g., the Konzo epidemic, discussed in Chapter 3). For terrestrial environments, exposure to anthropogenic sources of cyanide occurs primarily through contact with soils (Figure 17.3). Invertebrates living within the soil medium can come into direct contact with cyanide compounds. Plants can also be exposed via root systems. Exposure to wildlife species can occur via incidental ingestion of soils while foraging. Exposure to inorganic forms of cyanide via food chain pathways is not considered significant for wildlife species, as these compounds do not bioaccumulate [3]. It should be noted that some sources of solid waste containing cyanide in terrestrial ecosystems (e.g., spent potliner from aluminum manufacturing and oxide box waste from former MGP sites) also contain other chemicals (e.g., polycyclic aromatic hydrocarbons) that can present a hazard to terrestrial ecological receptors.

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17.1.1.2 Receptors The receptors that are most likely to be exposed to toxic forms and quantities of cyanide arising from anthropogenic inputs are those in intimate contact with environmental media (Figures 17.1 to 17.3). For aquatic systems, this includes invertebrates and fish that ventilate water over respiratory surfaces; for terrestrial systems, this includes soil invertebrates. While certain plants (e.g., cassava, blue-green algae) and animals (e.g., bacteria) can produce elevated levels of cyanide or cyanogenic compounds, this is primarily a natural process, as discussed in Chapters 3 and 12. Cyanide compounds derived from anthropogenic sources do not appear to undergo transfer or bioaccumulate in either aquatic or terrestrial food webs [3,15]. Incidental ingestion of soil and sediments containing cyanide could occur for wildlife that feed on soil/sediment invertebrates, and on plant roots and shoots. Beyer et al. [16] have estimated the percentage of sediment/soil in the diets of many wildlife species, including turtles, small mammals, and a number of aquatic and terrestrial bird species. From the normalized food ingestion rate (NFIR) and the percentage of sediment/soil in diet, the normalized sediment/soil ingestion rate can be determined. For mammals and birds that feed at the surface of sediments and soils, these percentages typically range between 5 and 20%, which is substantially higher than those associated with incidental ingestion of soils by humans (see discussion in Chapter 16). Therefore, wildlife species foraging on soil or sediment organisms receive greater doses (normalized to body weight) than do humans. Whether these exposures pose a risk depends upon the form and quantity of the cyanide ingested. However, this pathway is purely theoretical, as no data is available on the ingestion of cyanides from soil/sediment or soil/sediment organisms.

17.1.2 ASSESSMENT ENDPOINTS Assessment endpoints are explicit expressions of the target environmental parameter values that provide protection to ecological receptors. They are expressed in terms of an ecological entity (a species or type of ecological receptor) and an attribute (e.g., survival, growth, population sustainability). Based on experience, regulatory concerns, and the conceptual models illustrated in Figure 17.2 and Figure 17.3, certain assessment endpoints should be considered when cyanide is a chemical of potential concern. 17.1.2.1 Assessment Endpoints for Aquatic Environments The following assessment endpoints are suggested as starting points for ecological risk assessments involving release of cyanide in aquatic environments. They include: • • • • •

Ecological entities that are in direct contact with environmental media. Sustainability of planktonic invertebrate communities (that serve as a prey base). Sustainability of benthic invertebrate communities (that serve as a prey base). Sustainability of fish communities typical of the water body type. Survival of water-dependent wildlife that use the surface water for drinking purposes.

Each of these assessment endpoints combines an ecological entity such as “benthic invertebrate community” and an attribute such as “sustainability.” The word “sustainability” connotes the ability of the local population to sustain itself through reproductive replenishment. Thus, sustainability is usually judged through a combination of endpoints that are important for populations: the survival, growth, and reproduction of individual organisms that comprise the population. In some cases, sustainability is evaluated using population models that take measures of survival, growth, and reproduction as inputs. These assessment endpoints will not be appropriate for all cases, and there may be a need for additional assessment endpoints in some cases.

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17.1.2.2 Assessment Endpoints for Terrestrial Environments The following assessment endpoints are suggested as starting points for situations where surficial soils are contaminated with cyanide compounds. Presumably, such areas would have land uses commensurate with, and perhaps intended for, the support of ecological receptors. This may exclude land uses intended for industrial and commercial purposes. Wild and natural lands would typically be included, and agricultural, park, and residential lands may be included. Possible endpoints include: • • • •

Sustainability of soil invertebrate communities (for land uses intended as natural areas). Survival of ground-foraging mammals. Survival of ground-foraging birds. Sustainability of rooted terrestrial vegetation.

Again, these endpoints are suggested to aid in initiating discussions, but all of them may not be appropriate for all cases, and other assessment endpoints may be more important on a case-specific basis. The majority of the studies assessing cyanide toxicity in birds, wildlife, fish, and aquatic invertebrates were acute lethality tests resulting in oral LD50 or LC50 data [3]. No data are available for cyanide toxicity to soil invertebrates. Data on the chronic toxicity of cyanide to wildlife and aquatic species are limited (see Chapters 14 and 15).

17.2 ECOLOGICAL EFFECTS ASSESSMENT As explained in Chapters 13 to 15, free cyanide, comprising the cyanide ion CN− and molecular HCN, is the principal toxic form of cyanide for all receptors. The high affinity of CN− for many metals (e.g., Fe, Ni, Co, Cu, Zn, Ag, Au) results in the formation of metal–cyanide complexes. The chemistry of these metal–cyanide complexes is complicated, with great variations in the degree of affinity of the CN− for metals occurring in response to shifts in pH (Chapter 5). This characteristic affinity of CN− for metals is the basis for its many industrial applications, as well as its reactivity in biological systems. CN− also has a high affinity for sulfur, reacting with it to form thiocyanate, SCN− , a compound that has chemical properties similar to CN− , but is less toxic. This reduction in toxicity has led to the development of effluent treatment methods that produce SCN− as a means of CN− detoxification [17,18]. As explained in Chapters 7 and 8, free cyanide and metal–cyanide complexes are captured by the total cyanide measurement, which is often used to report the level of cyanide in water samples. Organocyanides are not captured in total cyanide measurements, but are themselves of little importance in aquatic toxicology.

17.2.1 METABOLIC BASIS OF CYANIDE REACTIVITY As a neutral hydrophilic molecule, HCN rapidly penetrates gill membranes from water. Inside the organism, HCN or its dissociated form CN− is transported to various target organs, where the strong nucleophilic center (CN− ) reacts readily with electron-deficient centers such as transition metals and sulfur moieties of various molecules; this essentially forms the basis of cyanide toxicity [19]. If CN− binds irreversibly with a sulfur atom from the sulfane–sulfur pool to form SCN− , it loses its high reactivity [20]. SCN− may also react with transition metals (Chapter 5). CN− attacks transition metal centers such as hemoglobin, cytochromes, and peroxidase enzymes, forming unstable complexes that will dissociate and be restored to “normal” function if the flow of CN− to the binding site is stopped. In contrast, when binding to a persulfide or sulfane–sulfur portion of a molecule occurs, the reaction is essentially irreversible. The many metabolically active enzymes and molecules that are affected by CN− and its mode of interference are thoroughly reviewed elsewhere [21,22].

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17.2.2 LETHALITY OF SCN−

AND

CN−

Exposure to acutely lethal concentrations of waterborne CN− in fish initially results in behavioral changes. There is an increase in ventilation rate, with fish surfacing and gulping air, and finally a frantic swimming, often in circles. Fish then lose control of swimming direction and buoyancy; this is often accompanied by convulsions and tremors. Finally, the fish die, with widely flared operculae [19]. Acute cyanide poisoning is a result of a sudden arrest of aerobic metabolism in the mitochondrial complexes of essential tissues, accompanied by the reduced oxygen carrying capacity of hemoglobin. Biochemically, CN− binds irreversibly to succinate dehydrogenase and reversibly to cytochrome a3 to disrupt the flow of electrons in the electron transport chain. CN− in the bloodstream inhibits catalase, an enzyme that breaks down metabolic hydrogen peroxide. Inhibition of catalase function allows the oxidation of iron in hemoglobin from Fe+2 to Fe+3 , reducing its affinity for oxygen and compounding the problem of arrest of aerobic metabolism by reducing the oxygen carrying capacity of hemoglobin [22]. Table 17.1 lists the single dose oral toxicity of free cyanide, added to water as the simple salt sodium cyanide, for various ecological receptors. A prominent feature of SCN− lethality is a group of clinical signs termed Sudden Death Syndrome (SDS) [23]. After the administration of a stressor (e.g., chasing with a dip net), SCN− exposed fish exhibit immediate loss of buoyancy, gasping, convulsions, rapid oscillating, pigmentation changes, extreme muscular rigor, flared operculae, and death within 20 sec. Many of the same signs occur during acute CN− poisoning, prompting Heming et al. [23] to suggest that one possible mechanism of death may be the result of a sudden shift in the physiological CN− /SCN− equilibrium towards the formation of CN− in vivo. However, it is not known whether the SCN− :HCN equilibrium ratio in fish blood is shifted during stress, or if endogenously generated CN− is actually involved in SDS. The complete clinical description of SDS has never been documented in acute CN− toxicity, although many of the signs are similar. SDS may be a direct effect of SCN− on neuromuscular functioning [23]. In summary, the lethal toxicity of CN− is manifested by a disruption of electron flow in the aerobic respiration pathway of fish, resulting in acute respiratory distress. The mechanisms involved in the acute toxicity of SCN− in fish have not been defined, but may include osmoregulatory and

TABLE 17.1 Single Dose Oral Toxicity of NaCN (mg/kg Body Weight) for various Ecological Receptors, Given as LD50 Values. (Species are Arranged from Most Sensitive to Most Tolerant) Oral LD50, NaCN (mg/kg bw, 95% CL)

Species Mallard (Anas platyrhynchos) American kesterel (Falco sparverius) Coyote (Canis latrans) Black vulture (Coragyps atratus) Norwegian rat (Rattus norvegicus) Little brown bat (Myotis lucifugus) Eastern screech owl (Otus asio) House mouse (Mus domesticus) Japanese quail (Coturnix japonica) European starling (Sturnus vulgaris) Domestic chicken (Gallus domesticus) White-footed mouse (Peromyscus leucopus)

2.7 (2.2–3.2) 4.0 (3.0–5.3) 4.1 (2.1–8.3) 4.8 (4.4–5.3) 5.1–6.4 8.4 (5.9–11.9) 8.6 (7.2–10.2) 8.7 (8.2–9.3) 9.4 (7.7–11.4) 17 (14–22) 21 (12–36) 28 (18–43)

Reference Henny et al. [6] Way [58] Sterner [59] Wiemeyer et al. [60] Ballantyne [61] Egekeze and Oehme [62] Clark et al. [63] Wiemeyer et al. [60] Clark et al. [63] Wiemeyer et al. [60] Wiemeyer et al. [60] Wiemeyer et al. [60] Clark et al. [63]

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nerve impulse transmission dysfunction. During acute poisoning by either chemical, many similar signs, such as flaring of operculae, loss of equilibrium, and erratic swimming, are evident.

17.2.3 SUBLETHAL TOXICITY Since CN− occurs naturally in the environment and is a metabolic by-product in a number of phyla, detoxification mechanisms have evolved to deal with low levels of CN− in body tissues. Smith et al. [24] have even suggested that CN− may actually be an essential trace metabolite. The most common form of CN− detoxification is the synthesis of SCN− from CN− and a labile sulfur atom catalyzed by rhodanese (thiosulfate–sulfur transferase), an enzyme found in the mitochondrial matrix [20]. CN− can also be reversibly immobilized by forming complexes with metals such as iron in hemoglobin or cobalt in Vitamin B12 and its precursors [24]. In this manner, the chronic toxicity of CN− is manifested in part by the biochemical actions of SCN− in tissues, and in part by the effects of CN− on sulfur- and metal-centered molecules. The most important effect of SCN− may be as an antithyroidal agent, inhibiting iodide uptake by the thyroid and the subsequent formation of thyroid hormones [25–28]. Lanno and Dixon [29–31] observed definite antithyroidal effects of SCN− in rainbow trout and fathead minnow. The most striking morphological effect of SCN− was the development of overt goitrous nodules in the branchial region, in both adult and juvenile fathead minnows. Histological examination confirmed the presence of a combination of diffuse hyperplastic and diffuse colloid goiter, which increased in severity with SCN− concentration. Goiter in rainbow trout was only detected histologically, but was similar in structure to goiter observed in fathead minnows. Previous observations on the antithyroidal effects of SCN− were derived almost entirely from mammalian research; the presence of an antithyroidal effect of waterborne SCN− in fish had not been previously verified [32]. Elevated plasma SCN− levels inhibit the movement of halides across membranes, such as those in the thyroid [28], the gills [33,34], and possibly the intestine [35]. CN− may also interfere with peroxidase activity in the synthesis of thyroxine from iodide and thyroglobulin. Due to the reduced availability of iodine, and possible interferences in the production of T3 and T4 , negative feedback on the pituitary results in enhanced production of Thyroid Stimulating Hormone (TSH). TSH acts upon the thyroid follicular cells to incorporate more iodine for thyroid hormone production. The net result is often seen as goiter, the hypertrophy or hyperplasia of the follicular cells of thyroid follicles [22,36]. Due to the metabolic demand on the pituitary for the synthesis of TSH, production may occur at the expense of Gonadotropic Hormone (GTH) synthesis, reducing reproductive efficiency [37]. Suggestions have been made that CN− may be regenerated from SCN− [38,39]. This has been observed to occur only in mammalian erythrocyte hemolysate preparations, and would appear to be an energetically and metabolically expensive “futile” cycle, requiring both energy and labile sulfur. It is also possible that the degradation of SCN− to CN− is an artefact of in vitro test systems, and may not occur in vivo.

17.2.4 COMPARATIVE SUBLETHAL TOXICITY OF CN−

AND

SCN−

Impairment of reproduction is the most prominent sublethal effect of chronic CN− exposure [40]. Fathead minnows were exposed to various concentrations of CN− for 265 d; the most sensitive toxicity endpoint was fecundity, with egg production per female decreasing by an average of 59% at 0.019 mg CN− /l [40]. Brook trout exposed to 0.011 mg CN− /l spawned fewer eggs than the control fish or even trout exposed to 0.006 mg CN− /l [41]. Above 0.054 mg CN− /l, no fertile eggs were spawned, with no eggs being released at 0.075 mg CN− /l. With rainbow trout, both altered patterns and incomplete deposition of secondary yolk were observed when maturing females were exposed to 0.01 or 0.02 mg CN− /l for 20 days, resulting in egg development being arrested with subsequent atresia of oocytes [42].

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Da Costa and Ruby [43] investigated the effect of CN− on serum calcium and phosphoprotein levels and hepatosomatic index (HSI) in both male and female rainbow trout. Exposure at 0.01 mg CN− /l resulted in reduced serum calcium levels relative to control values in both sexes, but serum calcium levels were not significantly different from control values at 0.02 and 0.03 mg CN− /l, respectively. CN− exposure for 7 d had no effect on HSI or serum phosphoprotein levels. Exposure of rainbow trout during early vitellogenesis to 0.01 mg CN− /l resulted in a significant depression in plasma vitellogenin levels compared to control vitellogenic fish at 12 d, but not at 6 d [44]. Gonadosomatic index (GSI) declined steadily with time in CN− -exposed fish, being significantly lower by 6 d, but HSI was not affected. Exposure of Atlantic salmon to 0.005 mg CN− /l for 12 d during late vitellogenesis resulted in a significant increase in plasma vitellogenin levels relative to control fish, but the vitellogenin content of ovarian and hepatic extracts was not affected [45]. Investigations on the effects of SCN− on reproduction in teleosts have been limited to studies where SCN− was administered by intraperitoneal injection. Singh et al. [37] administered 0.14 mg CN− /g fish/week for 8 weeks. This resulted in a decrease in I131 uptake by the thyroid (45% of control), decreased ovarian uptake of P32 (5% of control), and a decrease in gonadosomatic index (47% relative to controls). Although the number of eggs produced by each female was never quantified, fecundity may also be decreased by SCN− . The one problem with this study is that although biochemical parameters of reproductive status were measured, actual reproductive success was never monitored. In summary, the sublethal toxicity of CN− to fish is believed to be primarily a result of the metabolic conversion of CN− to SCN− , with the subsequent toxicity due to the action of SCN− [19]. Thiocyanate acts primarily as an antithyroidal agent, but may also inhibit the movement of various halides (Cl− , I− ) and phosphorus across membranes. It is the antithyroidal effect of SCN− that is thought to result in the impairment of reproduction in fish exposed to sublethal levels of CN− or SCN− .

17.2.5 SPECIES SENSITIVITY DISTRIBUTIONS Ambient Water Quality Criteria (AWQC), as well as site-specific risk assessments, consider information on the relative sensitivity of species to chemicals. For example, for AWQC, it is common to develop criteria that protect a percentage (approximately 95%) of the species that may occur in aquatic environments (see Chapter 14). Site-specific considerations take into account the types of species that a water body would support (e.g., warmwater fish or coldwater fish), as well as other factors that might influence exposure (primarily, bioavailability of the chemicals). The sensitivity of fish and aquatic invertebrates has been widely studied in laboratory exposures, with more tests conducted with fish, and more acute data than chronic data available. The acute toxicity of dissolved free cyanide varies with species as well as with duration of exposure. Figure 17.4 illustrates the relative sensitivities of aquatic animals to acute exposures of free cyanide. Coldwater fish species such as trout are the most sensitive animals tested, while aquatic invertebrates are less sensitive. Warmwater fish species such as bluegill sunfish are intermediate in sensitivity. Since decreased oxygen levels appear to exacerbate cyanide toxicity [3], the relative sensitivities of the species depicted in Figure 17.5 may reflect relative sensitivity to asphyxiation. Coldwater fish species such as trout may be more sensitive to oxygen deprivation than warmwater fish or aquatic invertebrates. Species sensitivity distributions for aquatic organisms exposed to free cyanide were developed from the toxicity data presented in Chapter 14. Only acute toxicity data meeting the classification criteria of Stephan et al. [46], which relate to acceptable experimental protocols and data analysis techniques, were used in the development of a species sensitivity distribution. Studies meeting these criteria were available for a number of aquatic organisms, including freshwater and marine fish, decapods, mysids, isopods, mollusks, and insects. Overall, acute toxicity data were available for 28 freshwater species, with species mean acute values (SMAVs) ranging from 46 to 10,000 µg CN/l.

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Acute LC50 (mg/l)

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100 CW fish WW fish Invertebrates

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1 Toxicity studies

FIGURE 17.4 Relative sensitivity of aquatic species to acute exposure of free cyanide.

Acute toxicity data were available for only 14 saltwater species, with SMAVs ranging from 59 to >10,000 µg CN/l, similar to freshwater species. Chronic data were available for only five freshwater and two saltwater species, with chronic values ranging from 8 to 79 µg CN/l. Further insight into the nature of species sensitivity distributions can be gained from constructing probability plots of sensitivity to free cyanide. One of the challenges for establishing these distributions is the variety of methods used to measure cyanide in the exposure medium, and the extent to which such variety obscures the true measure of cyanide exposure. The acute and chronic cyanide toxicity data for freshwater and saltwater species are summarized in Tables 14.1 to 14.5. In addition to these data meeting criteria for AWQC development [46], the data have been examined based on the analytical methods used to measure cyanide. The attention to analytical methods in Chapter 14 is illustrative of the complexity of determining cyanide exposure in empirical laboratory toxicity tests. Although it is accepted that total cyanide levels may not be well correlated with observed toxicity, there is not a single method that has historically been widely accepted as the correct analytical method for determining free cyanide (see Chapter 7). For the data evaluation conducted in Chapter 14, studies in which free cyanide was measured in some way were considered to offer the most accurate measure of exposure. However, restricting measurements of cyanide exposure to only these groups reduces the number of data points available for the estimation of a species mean acute value (SMAV). Examining the probability distributions of SMAVs for freshwater organisms, it appears that whether cyanide exposure is expressed as free cyanide, or whether any other measurement methodology is used, this does not drastically affect the distribution of SMAVs (Figure 17.5). Figure 17.5(a) includes all acute toxicity data, regardless of whether or how cyanide was measured (50th percentile SMAV of 158 µg CN/l); Figure 17.5(b) only includes data from studies where free cyanide was measured by a specified method (all species, 100 µg CN/l, 50th percentile SMAV); and Figure 17.5(c) is similar to Figure 17.5(b), but only includes data from studies conducted with freshwater fish (100 µg CN/l, 50th percentile SMAV). The 5th percentile SMAV ranges from 45 to 56 µg/l, regardless of which distribution is used to estimate it, and is driven by coldwater fish SMAVs (also see Chapter 14). The use of species sensitivity distributions for site-specific risk assessment will depend greatly on the nature of the water body. One possible application involves an assessment of potential risks to the integrity of a water body as a result of unintentional releases. In such cases, the risk would depend on the types of biota that reside in the water body. If the water body supports cold water fish species, these would likely be at greater risk than if the water body was a warmwater system. Differences would also likely exist between freshwater and coastal or marine systems. Another application of information on species sensitivity distributions could involve cost effectiveness evaluations of alternative remedial or treatment options. In such cases, a problem exists, and judgments are being made about alternatives that have the greatest potential for reducing the

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FIGURE 17.5 Probability distributions of SMAVs for acute cyanide toxicity in freshwater organisms (data from Chapter 14). (a) All acute toxicity data for all species, regardless of whether or how cyanide was measured; (b) only includes data from studies where free cyanide was measured by a specified method; (c) only includes data from studies conducted with freshwater fish, where free cyanide was measured by a specified method.

risks for a commensurate cost. Because risk is related to the sensitivity of species, and because water bodies vary in the types of species naturally present, some insights can be gained from understanding how different remedial or treatment technologies would affect (or not materially affect) the biota resident in receiving waters. The amount of risk reduction achievable with a particular technology will need to be balanced with the cost of installing and operating the technology. Issues of technology selection for cyanide management in different situations are discussed in Chapters 25 to 27. A third application of this information is in Stressor Identification (SI). SI tools are used to diagnose the chemical, physical, or biological stressors that result in water quality impairments [47]. Total Maximum Daily Load (TMDL) programs drive much of the SI effort [48]. However, the methodology is also being considered for other applications where waters are already impaired. This can include the evaluation of whether discharges to aquatic environments from various hazardous waste sites are causing impairments. SI tools are especially helpful where there are multiple types of stressors. Information on species sensitivity to cyanide can be used to examine whether or not those more sensitive species are absent from areas of potential exposure when they are present in suitable

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reference areas. These types of evaluations are very useful and are widely practiced. However, confounding factors are the presence of multiple chemicals in an effluent containing cyanide, which is generally the case, and the issue of whether the biological species of interest are equally sensitive to the mixture of chemicals. No data are available on the toxicity of cyanide in mixtures with other chemicals, so it would be difficult to separate the effects of a particular chemical such as cyanide in a mixture. When these types of mixture situations arise, it is important to determine which chemical(s) is (are) causing toxicity. Toxicity Identification Evaluations (TIEs) can be carried out for water or sediments to determine the relative toxicity of chemicals in the various fractions of the contaminated medium [49,50].

17.3 ECOLOGICAL EXPOSURE ASSESSMENT 17.3.1 AQUATIC ENVIRONMENTS Toxic forms of cyanide can be persistent in groundwater, and are present in certain wastewater effluents. Once discharged to surface water, cyanide can react with environmental constituents such as metals, sulfur-containing compounds, and other ligands to reduce toxicity, with the duration of toxicity dependent upon the cyanide loading in the discharge (Chapter 9). Few data are available on the toxicity of metal–cyanide complexes and other cyanide molecules. As described earlier, the toxicity of cyanide in aquatic environments is usually considered with respect to exposure to free cyanide. The trophic transfer of cyanide within aquatic systems is generally limited to naturally produced cyanogenic glycosides in plant material. Inorganic cyanide compounds — such as those associated with industrial activities — do not bioaccumulate in animal or plant tissues in aquatic environments to any significant degree [3]. Therefore, the principal exposure pathway for inorganic cyanide in aquatic ecosystems is via the water column. Exposure levels for inorganic cyanides in aquatic environments can be determined through free cyanide measurements (see Chapter 7) or through application of appropriate fate models (Chapters 9 and 10). The rationale for the water column as the major exposure pathway is that significant cyanide mass enters aquatic environments, primarily from industrial effluents, as a continuous discharge or as a pulse-exposure release due to pipeline breakages and other accidental releases. The cyanide concentrations of the sources can vary considerably, and these starting concentrations and types of releases for cyanide dictate whether ecologically significant acute or chronic effects will result. As noted earlier, total cyanide concentrations in wastewaters and leachates can range from ppb levels up to thousands of ppm. Discharges to surface waters can have significant impacts. For example, the occurrence of acutely toxic events (fish kills) in receiving waters has been associated with failures of containment systems for mining leach waters ([3]; and Chapters 1 and 26). Accidental pulse releases can be carried downstream, and remain at acutely toxic levels, depending on the degree of dilution afforded by the water body and the rates of other processes that reduce the levels of free cyanide in the receiving waters. Figure 17.6 provides a conceptual view of what occurs when a cyanide solution is introduced to a stream as a pulse release. Concentrations in the receiving water are reduced by dilution, as well as by other fate and transport processes described in Chapters 9 to 11. The effects of the release on downstream biota will depend on the concentrations of cyanide to which they are exposed, as well as to the duration of exposure. Acutely toxic exposures can also occur for wildlife that drink surface waters containing cyanide solutions (e.g., from wastewater impoundments). This exposure does not require that the cyanides be immediately available in the free dissolved phase. Cyanides that fall into the weak-acid-dissociable category may become liberated within the digestive system of birds and mammals [7,51]. While accidental pulse releases of cyanide solutions to surface waters have resulted in obvious short-term toxic events, chronic releases of lower-concentration solutions can result in longerterm exposures. Two common situations where chronic exposures occur are discharge of industrial wastewater effluents and discharge of groundwater that is contaminated with cyanides. The latter

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FIGURE 17.6 Conceptual view of changes in cyanide exposure following a pulse release to a stream.

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FIGURE 17.7 Conceptual view of changes in cyanide exposure following a continuous release to a stream.

typically results from leaching of waste materials (e.g., leaching from spent carbon potliner associated with aluminum manufacture, leach ponds from the mining industry, leachate from oxide box wastes at MGP sites) or by inadvertent or intended releases of wastewater effluents to the subsurface. Exposure concentrations from continuous sources typically vary with time, reflecting changes in the discharged or released quantities of cyanide, as well as temporal variations in groundwater or surface water hydrology. Because the toxicity of cyanides is influenced by duration of exposure as well as magnitude [3], and because organisms can recover from short-term exposures to low concentrations of dissolved cyanide, accurate estimates of risk are difficult to make when the exposure regime is temporally variable. Once released to surface waters, cyanides delivered through an outfall or from groundwater are subject to the fate processes described in Chapters 9 to 11. Figure 17.7 illustrates the changes in concentrations of free cyanide that arise from changes in distance from a chronic source. At actual sites, this spatial pattern of exposure will vary over time in accordance with load levels, changes in surface water hydrology, and the rates of fate processes (e.g., volatilization, metabolism). For risk assessment purposes, it is common to reflect exposures in terms of central tendency and upper bound values. Since fate processes occur at various time scales (e.g., diurnal, seasonal) and because cyanides can act at both short- and long-term time scales, data or estimates of exposure are best expressed at these various time scales. The analyses of exposure to chronic releases can also be expressed using probabilistic methods, such as joint probability analyses [52].

17.3.2 TERRESTRIAL ENVIRONMENTS Cyanide seldom remains biologically active (i.e., as free cyanide) in soils, since it is either complexed with metals (primarily Fe and Mn), or metabolized by microbes, or lost to soil air and thence to

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surface air through volatilization [53–55]. At waste deposits, leachable cyanide can be present, and can be a source of exposure to soil invertebrates and plants [56]. Wildlife that forage in these areas can also be exposed to cyanides by incidentally ingesting impacted soils or waste material [7,51]. For wildlife species with acidic digestive systems, some cyanide compounds can be converted to toxic forms and be absorbed within the gut. It has been proposed that cyanides that fall into the weak-acid-dissociable category can become liberated within the digestive system of birds and mammals [51]. Extractions of soils and wastes using approaches that simulate the cyanides that are liberated in the gut have been used as the basis for evaluating exposures via incidental soil or sediment ingestion. For example, the Massachusetts Department of Environmental Protection’s physiologically-available cyanide method [57] attempts to measure “biologically available cyanide,” following an acid extraction (to simulate gastric juice) at elevated temperatures well above those that occur in the digestive system of mammals. This test, however, has not yet been accepted by USEPA, or voluntary consensus organizations like ASTM or APHA, for regulatory compliance testing. While it is possible that toxicity to wildlife can occur if the animals ingest material from waste deposits or soils containing elevated levels of cyanide, other factors limit the likelihood of such exposures. Deposit of wastes and soils with cyanide levels high enough to be potentially toxic to wildlife often limit the development of plant and invertebrate animal communities that would be sought as food by wildlife. The development of plant and invertebrate communities can be limited by physical factors (compacted nature of some soils and materials), lack of plant nutrients, pH (the low pH of oxide box wastes and the high pH of spent potliner), and toxicity. As a result, locations where cyanide wastes are disposed on the land surface typically do not offer good forage for wildlife species. Thus, there is a low potential for incidental ingestion of soils that would be associated with foraging behavior in these areas.

17.4 RISK CHARACTERIZATION Risk characterization involves merging information on exposure with information on toxic effects. For cyanide exposures, risks may be characterized with respect to short (acute) exposure events, such as those involving pulse releases to aquatic environments, or situations where wildlife might drink or ingest water or other substances with high cyanide content. Risks may also be characterized in terms of chronic (long-term) exposures. In each case, it is critically important to match the toxicological data with the appropriate exposure duration (e.g., acute or chronic). For example, if there are short-term occurrences of elevated cyanide, it is most appropriate to match such intermittent exposures to toxicity data derived from acute (short-term) exposures, and not to data derived from long-term exposures (see related discussion in Chapter 16). The proper matching of exposure and toxicity data is probably one of the largest challenges in characterizing risks associated with cyanide exposures. It is also critically important, as suggested above, to express cyanide exposures in the appropriate dose metrics. Ideally, these would be some measures of bioavailable cyanide. Cyanide risks to aquatic systems depend on the magnitude and duration of exposures. The most challenging aspect of evaluating the ecological risk of inorganic cyanides is quantifying the exposure. This stems from the fact that the most toxic forms of cyanide (especially HCN) are short-lived, while less toxic forms, such as metabolites (e.g., SCN− ), are persistent. Assessing exposure has also been confounded by analytical methods employed (e.g., free cyanide versus total cyanide), some of which do not easily translate into metrics that can be used to judge exposure in a meaningful way. We believe that because “cyanide” draws such attention as a chemical compound, risk practitioners and regulators must exercise considerable care when evaluating exposures and risk. It is very tempting to make simple assumptions on exposure and toxicity that result in very erroneous risk estimates. We also note that while there is considerable information on acute effects, there are relatively few data on chronic effects.

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17.5 SUMMARY AND CONCLUSIONS • Free cyanide is the principal toxic form of cyanide for ecological recetors. • The most recognized effects of cyanide releases on aquatic systems have been associated with sudden unintentional releases of large quantities (slug or pulse releases), with risk dependent upon the volume of the receiving water body, as well as on the nature of the biota. • Chronic exposure to free cyanide in surface waters is an important consideration for cases in which wastewater effluents and groundwater discharges are sources, though discharged cyanides are subject to various fate processes that act to reduce concentrations. • Inorganic sources of cyanide do not bioaccumulate in aquatic or marine food webs, limiting exposure through trophic transfer. • The most complete aquatic toxicity data set available is for acute toxicity in freshwater organisms. Risk characterization and the development of water quality criteria are driven by coldwater fish data (e.g., rainbow trout). • Ingestion of surface waters from treatment or storage impoundments containing cyanide is a source of cyanide exposure to birds and mammals, and can result in acute and chronic toxic effects. • Risks to wildlife that forage near areas where solid wastes are disposed will depend mainly on the degree to which they ingest soils that contain cyanides, and whether such areas actually support food items such as plants and invertebrates on which the wildlife feed. • Acute cyanide toxicity to wildlife species is well known through cyanide use for pest control (e.g., for prairie dogs and coyotes). However, these forms of cyanide differ greatly from the dissolved metal–cyanide complexes and cyanide-bearing solid wastes associated with industrial wastes. • The most challenging aspect of evaluating the ecological risk of inorganic cyanides is quantifying the exposure. This stems from the fact that the most toxic forms of cyanide (especially HCN) are short-lived, while less toxic forms, such as metabolites (e.g., SCN− ), are persistent.

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10. Rubec, P.J., The need for conservation and management of Phillipine coral reefs, Environ. Biol. Fish., 23, 141, 1988. 11. Johannes, R.E. and Riepen, M., Environmental, economic, and social implications of the live reef fish trade in Asia and the western Pacific, The Nature Conservancy, Jakarta, Indonesia, 1995. 12. Barber, C.V. and Pratt, V.R., Sullied seas: strategies for combating cyanide fishing in Southeast Asia and beyond, World Resources Institute, Washington, DC, 1997. 13. Barber, C.V. and Pratt, V.R., Poison for profits: cyanide fishing in the Indo-Pacific, Environment, 40, 5, 1998. 14. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sediment Contam., submitted, 2005. 15. USEPA, Ambient water quality criteria for cyanide-1984, EPA-440/5-84-028, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 1984. 16. Beyer, W.N., Connor, E.E., and Gerould, S., Estimates of soil ingestion by wildlife, J. Wildlife Manage., 58, 375, 1994. 17. Devuyst, E.A., Conrad, B.R., and Ettel, V.A., Pilot plant operation of the Inco SO2 /air cyanide removal process, Can. Mining J., August 27, 1982. 18. Ingles, J. and Scott, J.S., State-of-the-art processes for treatment of gold mill effluents, Mining, Mineral and Metallurgical Processes Division, Environment Canada, Ottawa, Ontario, 1987. 19. Leduc, G., Cyanides in water: toxicological significance, in Aquatic Toxicology, Vol. 2, Weber, L.J., Ed., Raven Press, New York, 1984, p. 153. 20. Westley, J., Cyanide and sulfane sulfur, in Cyanide in Biology, Vennesland, C.B., Conn, E.E., Knowles, C.J., and Wissing, F., Eds., Academic Press, New York, 1981, p. 61. 21. Vennesland, B., Castric, P.A., Conn, E.E., Solomonson, L.P., Volini, M., and Westley, J., Cyanide metabolism, Fed. Proc., 41, 2639, 1982. 22. Leduc, G., Pierce, R.C., and McCracken, J.R., The effects of cyanides on aquatic organisms with emphasis upon freshwater fishes, National Research Council of Canada, Ottawa, Ontario, 1982. 23. Heming, T.A., Thurston, R.V., Meyn, E.L., and Zajdel, R.K., Acute toxicity of thiocyanate to trout, Trans. Am. Fish. Soc., 114, 895, 1985. 24. Smith, A.D.M., Duckett, S., and Waters, A.H., Neuropathological changes in chronic cyanide intoxication, Nature, 200, 179, 1963. 25. Green, W.L., Antithyroid compounds, in The Thyroid: A Fundamental and Clinical Text, Ingbar, S.H. and Braverman, L.E., Eds., J.B. Lippincott Company, Philadelphia, PA, 1986, p. 339. 26. Yamada, T., Kajihara, A., Takemura, Y., and Onaya, T., Antithyroid compounds, in Handbook of Physiology, Section 7: Endocrinology, Vol. III. Thyroid, Greep, R.O. and Astwood, E.B., Eds., American Physiological Society, Washington, DC, 1974, p. 345. 27. Wood, J.L., Biochemistry, in Chemistry and Biochemistry of Thiocyanic Acid and Its Derivatives, Newman, A.A., Ed., Academic Press, New York, 1975, p. 156. 28. Wolff, J., Transport of iodide and other anions in the thyroid gland, Physiol. Rev., 44, 45, 1964. 29. Lanno, R.P. and Dixon, D.G., Chronic toxicity of waterborne thiocyanate to rainbow trout (Oncorhynchus mykiss), Can. J. Fish. Aquat. Sci., 53, 2137, 1996. 30. Lanno, R.P. and Dixon, D.G., Chronic toxicity of waterborne thiocyanate to the fathead minnow (Pimephales promelas): a partial life-cycle study, Environ. Toxicol. Chem., 13, 1423, 1994. 31. Lanno, R.P. and Dixon, D.G., The comparative chronic toxicity of thiocyanate and cyanide to rainbow trout (Oncorhynchus mykiss), Aquat. Toxicol., 36, 177, 1996. 32. Eales, J.G. and Shostak, S., Influence of potassium thiocyanate on thyroid function of rainbow trout, Salmo gairdneri, Gen. Comp. Endocrinol., 51, 39, 1983. 33. Maetz, J. Transport of ions and water across the epithelium of fish gills, in Proceedings of the Symposium on Lung Liquids, CIBA Foundation Series No. 38, London, 1975, p. 133. 34. Epstein, F.H., Silva, P., Forrest, J.N., and Solomon, R.J., Chloride transport and its inhibition by thiocyanate in gills of seawater teleosts, Comp. Biochem. Physiol., 52A, 681, 1975. 35. Katz, U., Lau, K.R., Ramos, M.M.P., and Ellory, J.C., Thiocyanate transport across fish intestine (Pleuronectes platessa), J. Membrane Biol., 66, 9, 1982. 36. Beckers, C., Nontoxic goiter, in The Thyroid Gland, DeVissher, M., Ed., Raven Press, New York, 1980, p. 234.

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37. Singh, R., Raizada, R.B., and Singh, T.P., Effects of some antithyroid drugs on the pituitary–thyroid– gonad axis in a freshwater catfish, Heteropneustes fossilis (Bloch), Gen. Comp. Endocrinol., 31, 451, 1977. 38. Chung, J. and Wood, J.L., Oxidation of thiocyanate to cyanide catalyzed by hemoglobin, J. Biol. Chem., 246, 555, 1971. 39. Goldstein, F. and Reiders, F., Conversion of thiocyanate to cyanide by an erythrocytic enzyme, Am. J. Physiol., 173, 287, 1953. 40. Lind, D.T., Smith, L.L., and Broderius, S.J., Chronic effects of hydrogen cyanide on the fathead minnow, J. Water Poll. Control Fed., 49, 262, 1977. 41. Koenst, W.M., Smith, L.L., and Broderius, S.J., Effect of chronic exposure of brook trout to sublethal concentrations of hydrogen cyanide, Environ. Sci. Technol., 11, 883, 1977. 42. Lesniak, J.A. and Ruby, S.M., Histological and quantitative effects of sublethal cyanide exposure on oocyte development in rainbow trout, Arch. Environ. Contam. Toxicol., 11, 343, 1982. 43. DaCosta, H. and Ruby, S.M., The effect of sublethal cyanide on vitellogenic parameters in rainbow trout Salmo gairdneri, Arch. Environ. Contam. Toxicol., 13, 101, 1984. 44. Ruby, S.M., Idler, D.R., and So, Y.P., The effect of sublethal cyanide exposure on plasma vitellogenin levels in rainbow trout (Salmo gairdneri) during early vitellogenesis, Arch. Environ. Contam. Toxicol., 15, 603, 1986. 45. Ruby, S.M., Idler, D.R., and So, Y.P., Changes in plasma, liver, and ovary vitellogenin in landlocked Atlantic salmon following exposure to sublethal cyanide, Arch. Environ. Contam. Toxicol., 16, 507, 1987. 46. Stephan, C.E., Mount, D.I., Hansen, D.J., Gentile, J.H., Chapman, G.A., and Brungs, W.A., Guidelines for deriving numerical national water quality criteria for the protection of aquatic organisms and their uses, EPA-600/4-85-014, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1985. 47. USEPA, Stressor identification guidance document, EPA-822-B-00-025, U.S. Environmental Protection Agency, Washington, DC, 2000. 48. Cormier, S., U.S. Environmental Protection Agency, Washington, DC, personal communication, 2005. 49. USEPA, Methods for aquatic toxicity identification evaluations: Phase I toxicity characterization procedures, EPA-600/3-88-034, U.S. Environmental Protection Agency, Duluth Environmental Research Laboratory, Duluth, MN, 1991. 50. USEPA, Sediment toxicity identification evaluation: Phase I (characterization), Phase II (identification), and Phase III (confirmation) modifications of effluent procedures, EPA-600/6-91-007, U.S. Environmental Protection Agency, Duluth Environmental Research Laboratory, Duluth, MN, 1991. 51. Clark, D.R. and Hothem, R.L., Mammal mortality at Arizona, California, and Nevada gold mines using cyanide extraction, California Fish Game, 77, 61, 1991. 52. Suter, G.W., Efroymson, R.A., Sample, B.E., and Jones, D.S., Ecological Risk Assessment for Contaminated Sites, Lewis Publishers, Boca Raton, FL, 2000. 53. Köster, H.W., Risk assessment of historical soil contamination with cyanides: origin, potential human exposure and evaluation of intervention values, RIVM report 711701019, Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment), Bilthoven, The Netherlands, 2001. 54. Towill, L.E., Drury, J.S., Whitfield, B.L., Lewis, E.B., Galyan, E.L., and Hammons, A.S., Review of the environmental effects of pollutants. V. Cyanide, EPA-600/1-78-027, U.S. Environmental Protection Agency, Office of Research and Development. Cincinnati, OH, 1978. 55. Marrs, T.C. and Ballantyne, B., Clinical and experimental toxicology of cyanides: an overview, in Clinical and Experimental Toxicology of Cyanides, Ballantyne, B. and Marrs, T.C., Eds., Wright Publishers, Bristol, U.K., 1987, p. 473. 56. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Clarification: equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 501, 1999. 57. MADEP, Quality assurance and quality control requirements and performance standards for SW 846 Method 9014, total cyanide and the MADEP physiologically available cyanide (PAC) protocol, WSC-CAM-VI A, Massachusetts Department of Environmental Protection, Bureau of Waste Site Cleanup, Boston, MA, 2004.

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58. Way, J.L., Pharmacologic aspects of cyanide and its antagonism, in Cyanide in Biology, Vennesland, C.B., Conn, E.E., Knowles, C.J., and Wissing, F., Eds., Academic Press, New York, 1981, p. 29. 59. Sterner, R.T., Effects of sodium cyanide and diphacinone in coyotes (Canis latrans): applications as predacides in livestock toxic collars, Bull. Environ. Contam. Toxicol., 23, 211, 1979. 60. Wiemeyer, S.N., Hill, E.F., Carpenter, J.W., and Krynitsky, A.J., Acute oral toxicity of sodium cyanide in birds, J. Wildlife Dis., 22, 538, 1986. 61. Ballantyne, B., Toxicology of cyanides, in Clinical and Experimental Toxicology of Cyanides, Ballantyne, B. and Marrs, T.C., Eds., Wright Publishers, Bristol, U.K., 1987, p. 41. 62. Egekeze, J.O. and Oehme, F.W., Cyanides and their toxicity: a literature review, Veterin. Quart., 2, 104, 1980. 63. Clark, D.R., Hill, E.F., and Henry, P.F.P., Comparative sensitivity of little brown bats (Myotis lucifugus) to acute dosages of sodium cyanide, Bat Res. News, 32, 68, 1991.

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of Cyanide in Water 18 Regulation and Soil David V. Nakles, David A. Dzombak, Rajat S. Ghosh, George M. Wong-Chong, and Thomas L. Theis CONTENTS 18.1

U.S. Regulations, Guidelines and Criteria for Cyanide in Water . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.1 Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.2 Surface Water and Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.3 Federal Wastewater Discharge Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.3.1 Effluent Guidelines and Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.3.2 Specific Characteristics of Guidelines/Standards. . . . . . . . . . . . . . . . . . . . 18.1.4 Groundwater (RCRA and CERCLA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.4.1 Groundwater (RCRA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.1.4.2 Groundwater (CERCLA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2 U.S. Regulations, Guidelines, and Criteria for Soil, Sediment, and Process Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.1 Soil and Process Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.1.1 Listed Wastes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.1.2 Characteristic Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.2 Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3 International Regulatory Standards and Guidelines for Cyanide in Water and Soil . . . . . 18.3.1 Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.1.1 Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.1.2 Surface Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.1.3 Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.2 Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.3 Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4 Technical/Regulatory Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.1 Lack of Consistency: Analytical Methods and Regulations. . . . . . . . . . . . . . . . . . . . . 18.4.2 Cyanide Transformation in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.3 Cyanide Toxicological Database . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.4 Cyanide as a CERCLA Hazardous Substance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

352 355 359 360 360 361 362 371 372 372 372 373 373 373 375 376 376 376 377 378 378 380 380 381 381 382 382 383

Various forms of cyanide in water and soil have been regulated in the United States and elsewhere for many years, dating back to the beginning of the environmental era of the early 1970s. The primary driver for regulating cyanide, of all forms, is the acute human and ecological toxicity associated with hydrogen cyanide. The toxicological effects of this compound on humans and animals have been extensively examined and are well understood (see Chapters 13–15). However, it is also understood 351

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that cyanide may exist in a number of different chemical forms, none of which are as toxic as hydrogen cyanide. In fact, several forms of cyanide are known to exist naturally in the environment and to serve as a source of nitrogen in the natural life cycle of plants and other organisms (Chapters 3 and 12). Analytical methods for the detection of these different forms of cyanide have been developed or approved for regulatory use relatively recently. These circumstances have led to fairly conservative cyanide regulations for soil and water that are in many cases based on the concentration of total cyanide. That is, many regulations do not discriminate among the various forms of cyanide that may be present. This was a reasonable approach for regulating cyanide-impacted media in the 1970s, given the previous state of the analytical methods and the potential for specific nontoxic forms of cyanide to release hydrogen cyanide under certain environmental conditions. However, as the science for the detection of cyanide species and the understanding of the fate of these species in the environment has evolved, modifications to the regulatory framework have begun to be implemented at both the state and Federal levels in the United States, and in other countries as well. This chapter presents a summary of the U.S. environmental regulations that address the forms of cyanide that may be present in a soil or water matrix. A brief examination of some water and soil regulations for cyanide in other countries is also provided. At the Federal level in the United States, cyanide in water and soil is regulated under the Clean Water Act (CWA), the Safe Drinking Water Act (SDWA), and the Resource Conservation and Recovery Act (RCRA). The regulations promulgated under these acts by the U.S. Environmental Protection Agency (USEPA) have set forth specific standards and criteria for cyanide in receiving water, drinking water, wastewaters, soil, and various wastes. Spills of regulated hazardous substances that contain cyanide may also invoke the requirements of Superfund, that is, the Comprehensive Environmental Response and Liability Act, or CERCLA. At the same time, there are number of state regulations that address cyanide, most of which have been derived from the existing Federal legislation. In addition to summarizing the regulations, this chapter also discusses a number of technical issues that can, and often do, complicate the strict application of the regulations. These issues can be grouped under the topics of cyanide speciation, cyanide analytical methods, environmental transformation of cyanide, and cyanide toxicology. Each of these topics is also discussed in more detail in other chapters of this book.

18.1 U.S. REGULATIONS, GUIDELINES AND CRITERIA FOR CYANIDE IN WATER Water quality-related standards and guidelines for cyanide in the United States are voluminous and complex. Tables 18.1–18.4 present a summary of this information according to the governing authority (national or state), regulatory focus (surface water, drinking water, wastewater), health effects target (human or aquatic life), cyanide form (total, free, complexed, amenable), if specified, and water usage or origin. State criteria are often assigned based on water usage category. Table 18.5 summarizes such categories as they are defined by selected states. A complete listing of all healthrelated cyanide standards as of 1997 can be found in the U.S. Department of Health and Human Services report on the Toxicological Profile for Cyanide [1]. A review of the information contained in Tables 18.1–18.4 reveals several features of interest. First, criteria for cyanide vary significantly depending on the designated use for the water. These criteria also reflect the large differential toxicities that cyanide compounds can exhibit among target organisms. For instance, numerical standards related to human consumption and exposure are generally considerably higher than those for sensitive aquatic organisms. Further, the most sensitive organism can vary depending on the specific aquatic environment (e.g., marine vs. fresh waters, or cold vs. warm waters). Second, in recognition of the toxicity differences among cyanide species (regardless of target organism), many criteria attempt to differentiate among chemical forms

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TABLE 18.1 National USEPA Guidelines for Cyanide Description 1-Day Health Advisory Child 10-Day Health Advisory Child Lifetime Health Advisory Adult Longer-term Health Advisory Child Adult Maximum contaminant level Copper cyanide, cyanide, potassium silver cyanide, sodium cyanide Maximum contaminant level goal Cyanide, potassium silver cyanide, sodium cyanide, potassium cyanide Copper cyanide Ambient water quality criteria for human health Potassium silver cyanide, sodium cyanide, potassium cyanide, copper cyanide (water and fish) Ambient water quality criteria for aquatic organisms Sodium cyanide, potassium cyanide Freshwater acute Freshwater chronic Marine acute Copper cyanide Freshwater acute Freshwater chronic Marine acute Proposed rule: Great Lakes system water quality standards Acute water quality criteria for protection of aquatic life Chronic water quality criteria for protection of aquatic life

Limit (µg/l)

Designation

220

Cyanide

220

Cyanide

200

Cyanide

200 800

Cyanide Cyanide

200

200 1300 200

22 5.2 1 9.2 6.5 2.9

22

Free cyanide

5.2

Free cyanide

Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

of cyanide, sometimes by specifically listing them (e.g., free cyanide, copper cyanide, potassium cyanide), or by specifying an operational class of cyanide compounds as measurable by an analytical procedure (e.g., cyanide amenable to chlorination). Third, and perhaps most critically, the degree of specificity among cyanide compounds and chemical forms listed among the various standards, when viewed collectively across national and state criteria, is inconsistent with available approved analytical methodologies; that is, many more chemical forms are recognized as being of importance than current approved analytical methods can accommodate.

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TABLE 18.2 State Human Health Standards for Cyanide State

Description

Arizona

Drinking water guideline Domestic water source (DWS) Fish consumption (FC) Full body contact (FBC) Partial body contact (PBC)

Colorado Connecticut

District of Columbia Florida

Iowa

Illinois Idaho Indiana Kansas Kentucky Massachusetts Maine Michigan Minnesota

New Hampshire

New Jersey New Mexico New York

North Carolina Ohio Oklahoma Oregon Rhode Island Tennessee Utah Virginia

Degree of treatment Disinfection and chemical Complete Maximum permissible level Class C Class D Domestic/drinking MCL (Maximum contaminant level) Criteria for surface waters, Class I–V MCL Class B waters Class C waters MCL MCL Continuous (4-day average) Point of water intake Drinking water guideline MCL: domestic water supply Drinking water guideline Drinking water guideline Domestic/drinking Drinking water guideline Class A and B waters Class D waters Drinking water guideline MCL Municipal/domestic Domestic/drinking Groundwater quality: Groundwater Domestic/drinking Groundwater Effluent Standards: maximum allowable concentration Surface waters and groundwater A, A-S, AA, AA-S GA Class GSA groundwater 30-day average Maximum allowable levels Domestic/drinking Drinking water guideline Domestic/drinking MCL Groundwater

Limit (µg/l)

Designation

220 140 210,000 3,100 3,100 200

Total cyanide Total cyanide Total cyanide Total cyanide

10 200 200 3 200 200 200 5.0 5 20 200 200 200 154 200 140 154 150 154 10 200 154 10 200 200 200 200 100 400

100 100 154 200 200 200 150 200 200 5

Free cyanide

Free cyanide (CN) (CN) (CN)

(CN)

Free cyanide

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TABLE 18.2 Continued State

Description

Vermont Wyoming Wisconsin

West Virginia

Drinking water standard Class A or B waters MCL — groundwaters Public water supplier Warmwater sport fish communities Cold water communities Great Lakes communities Nonpublic water supplier Warmwater sport fish communities Cold water communities Warmwater forage and limited forage Groundwater Enforcement standard Prevention action limit Water quality criteria Warm water fishery streams Trout waters Small nonfishable streams Water contact, recreation Water supply, public

Limit (µg/l)

Designation

154 200 200 600 600 600

Total cyanide Total cyanide Total cyanide

40,000 40,000 120,000

Total cyanide Total cyanide Total cyanide

200 40 5 5 5 5 5

Free cyanide Free cyanide Free cyanide Free cyanide Free cyanide

Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1987.

18.1.1 DRINKING WATER The national water quality criterion for the protection of human health, that is, the maximum contaminant level or MCL, developed by the USEPA, is 200 micrograms per liter (µg/l), measured as free cyanide (see Table 18.1). This value was developed under the SDWA; the same value is also designated the maximum contaminant level goal, or MCLG, for cyanide in drinking water. The concentration limit was derived based on the assumption that an adult can ingest 0.02 mg of cyanide per kilogram of body weight per day without causing an unacceptable adverse health effect. This acceptable dose was determined based on a study in which 10 female and 10 male rates were provided a range of cyanide doses (delivered as sodium cyanide) in drinking water for 13 weeks [1]. Note that copper cyanide has an MCLG (1300 µg/l) that is much higher than that of free cyanide. This elevated MCLG is indicative of the difference in toxicity between this form of cyanide and those typically included as part of the free cyanide, that is, hydrogen cyanide, potassium cyanide, potassium silver cyanide, and sodium cyanide. Table 18.1 also lists other health-related cyanide standards, all of which were abstracted from the U.S. Department of Health and Human Services report on the Toxicological Profile for Cyanide (Table 7.1, [1]). These standards are referenced as Health Advisories and include levels that are explicitly designated as applicable to either children or adults; however, none of these advisories designate the specific form of the cyanide to which they apply. The health advisories for children range from 200 (Longer-term Health Advisory) to 220 µg/l (1- and 10-day health advisories) of cyanide; only a Longer-term Health Advisory of 800 µg/l and a Lifetime Health Advisory of 200 µg/l are specified for adults.

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TABLE 18.3 State Aquatic Life Standards for Cyanide State Alabama

Arizona

Florida Hawaii

Indiana

Kentucky

Maryland

Minnesota Mississippi

Missouri

Nevada

New York

North Dakota North Carolina

Description

Limit (µg/l)

Freshwater:acute Freshwater: chronic Marine: acute Marine: chronic Acute criteria for aquatic & wildlife Cold water fishery (A&Wc) Warm water fishery (A&Ws) Effluent dominated water (A&Wedw) Ephemeral (A&We) Chronic criteria for aquatic & wildlife A&Wc A&Ws A&Wedw A&We Criteria for surface water, Class I-V Freshwater: acute (ecological standard) Freshwater: chronic (ecological standard) Saltwater: acute (ecological standard) Saltwater: chronic (ecological standard) Acute aquatic criterion Continuous (4-day average) outside of mixing zone: chronic aquatic criterion Maximum allowable instream conc Chronic (ecological standard) Acute (ecological standard) Ambient surface waters Freshwater: acute Freshwater: chronic Estuarine: acute Estuarine:chronic Saltwater: acute Saltwater: chronic Class A, B, C waters Freshwater: acute Freshwater: chronic Saltwater: acute Saltwater: chronic Chronic toxicity

22.0 5.2 — — 1.0 — 20.0 22.0 5.2 1.0 1.0 5.0

Acute toxicity

22.0

Single value 24-h average Propagation of wildlife Surface waters & groundwaters A, A-S, AA, AA-S, B, C D SA, SB, SC SD Class I streams Freshwater

52.0 3.5 5.0

Designation

22.0 5.2 1.0 — 22.0 41.0 41.0 84.0

T (total recoverable) T T T

5.2 9.7 9.7 19.0 5.0 22 5.2 1.0 1.0 22.0 5.2

T T T T

5 22

5.2 22.0 1.0 1.0 5.0 5.0

Free cyanide Free cyanide

(CN)

Amenable to chlorination Amenable to chlorination

Total cyanides

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TABLE 18.3 Continued State Puerto Rico Oklahoma Ohio

Virginia Vermont Wyoming Wisconsin

Description

Limit (µg/l)

Coastal estuarine waters Surface waters Acute Chronic Outside mixing zone (maximum) Cold water Limited resource warm water 30-day average cold water Inside mixing zone (maximum) Cold water Limited resource warm water Freshwater Saltwater Acute Chronic Special A waters Great Lakes Cold water Warm water sport fish All others

Designation

20.0 20.0 45.93 10.72 46.0 22.0 5.2 45.0 92.0 5.2 1.0 22.0 5.2 5.0 22.4 22.4 46.2 46.2

Free cyanide Free cyanide Free cyanide

Total cyanide Total cyanide

Free cyanide Free cyanide Free cyanide Free cyanide

Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

TABLE 18.4 Miscellaneous State Standards for Cyanide State Water quality: agricultural uses Arizona

Nevada Hazardous constituents Indiana

Description

Agricultural irrigation (AgI) Livestock watering (Ag L) Ag L

Allowable concentration using leaching test method: Class IV Class III Class II

Limit (µg/l)

No numerical standard 200

Designation

Total recoverable

200

200 2000 5000

Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

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TABLE 18.5 Selected State Water Classifications Florida Class I — Potable water supplies Class II — Shellfish propagation or harvesting Class III — Recreation, propagating, and maintenance of healthy, well-balanced population of fish and wildlife Class IV — Agricultural water supplies Class V — Navigation, utility, and industrial use Minnesota Class 1 — Domestic consumption Class 2 — Aquatic life and recreation New Jersey GW1 — Ground water of special ecological significance GW2 — Ground water for potable water supply GW3 — Ground water with uses other than potable water supply New York Class N — fresh surface waters Class A, AA, Special (AA-S) — fresh surface waters (drinking) Class B — fresh surface waters (primary and secondary contact) Class C, D — fresh surface waters (fishing) Class SA — saline surface waters (fish propagation and survival) Class SB — saline surface waters (primary and secondary contact) Class SC — saline surface waters (fishing) Class I — saline surface waters (secondary contact recreation) Class SD — saline surface waters (fish survival) Class GA — fresh groundwaters (drinking) Class GSA — saline groundwaters (potable mineral waters) Class GSB — saline groundwaters (receiving waters) North Carolina Class GA — groundwaters (drinking water) Class GSA — groundwaters; usage and occurrence (potable mineral waters) Class GC — groundwaters; usage and occurrence (nondrinking uses) Vermont Class A(1) — Ecological waters Class A(2) — Public water supplies Class B — Cold and warm water fish habitats Wyoming Special A waters — Suitable for fish and aquatic life Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

A review of the human health standards and guidelines for various states (Table 18.2) indicates a greater degree of variability in the concentrations of cyanide that are acceptable in drinking water. An examination of all 50 states reveals that most states do adopt, in some manner, the national MCL of 200 µg/l free cyanide. However, there are often variations from this value based on a specific classification of the water that is being consumed. For example, Arizona has a drinking water guideline of 220 µg/l free cyanide but also has a standard of 140 µg/l of total cyanide for domestic water sources (DWS). Similarly, Massachusetts and Maine have drinking water guidelines

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of 140 and 154 µg/l, respectively, although they do not indicate the form of cyanide that should be measured. New Hampshire has a similar drinking water guideline (i.e., 154 µg/l) but also has its own MCL of 10 µg/l, measured as [CN]. This value represents the lowest human-health water quality standard for cyanide that exists among all of the 50 states.

18.1.2 SURFACE WATER AND GROUNDWATER Section 304(a)(1) of the Clean Water Act requires USEPA to develop and publish ambient water quality criteria for selected pollutants, of which cyanide is one. Ambient surface water quality criteria for aquatic organisms are provided to protect against both chronic and acute toxicological effects. Simply stated, acute toxicity is toxicity that occurs very rapidly, for example, minutes to hours, after exposure whereas chronic toxicity occurs only after exposure over long periods of time, for example, days to years. The chronic and acute ambient water criteria for cyanide in freshwater are 5.2 and 22 µg/l, respectively, expressed as free cyanide (as CN). These criteria were developed based on trout toxicity data (Chapter 14). There is little debate about the validity of these freshwater criteria. However, the same is not true for saltwater criteria, which are 1.0 µg/l for both chronic and acute toxicity. These criteria are at or below the detection limit for available analytical methods (Chapter 7). Further, some have suggested that the test organism used by the USEPA to derive the saltwater criteria, Cancer irroratus, is uncommon and atypically sensitive [2]. This issue is examined in detail in Chapter 14, where the aquatic toxicity database underlying the national ambient water quality criteria for cyanide is summarized and critiqued. Surface water quality criteria are also defined specifically for copper cyanide with values of 9.2, 6.5, and 2.9 µg/l respectively, for acute and chronic effects in freshwater and both acute and chronic effects in saltwater (see Table 18.1). The Clean Water Act also directs USEPA to develop ambient water quality criteria for protection of human health, especially for exposure through consumption of fish and also including incidental water consumption, for example, through recreation activities. The criteria are developed using a methodology that incorporates a set of standard data and approaches for evaluating exposure and health risk [3]. The human health water quality criteria for cyanide were revised in 2003 [4]. The criterion for both “consumption of water and organism” and “consumption of organism only” is 140 µg/l total cyanide. It is noted in the USEPA criteria document [3] that the recommended water quality criterion is expressed as total cyanide even though the Integrated Risk Information System (IRIS) reference dose used to derive the criterion is based on free cyanide. USEPA notes that “if a substantial fraction of the cyanide present in a water body is present in complexed form (e.g., Fe4 [Fe(CN)6 ]3 ), this recommended criterion may be overly conservative.” State surface water criteria are often assigned based on water usage categories (Tables 18.3–18.5). Typical examples of these categories include potable water supplies, recreational use, agricultural water supplies, groundwater, saline/fresh surface water, and cold/warm water fish habitats, to name a few. A review of the state aquatic life protection criteria in Table 18.3 reveals a range of concentrations for total cyanide and total recoverable cyanide (1.0 µg/l [saltwater/acute and chronic effects (several states); New York: selected freshwater effects] to 84 µg/l [Arizona]); cyanide amenable to chlorination (Several states: 5 µg/l [chronic toxicity] to 22 µg/l [acute toxicity]), and free cyanide (5.2 µg/l [Ohio: 30-day average cold water] to 46.2 µg/l [Wisconsin: warm water sport fish and all others]). Cyanide water quality criteria for groundwater are not common. Those that do exist usually treat the groundwater as either a potential drinking water or as a potential source to an adjacent surface water. In the former instance, the criteria are usually based on the MCL or 200 µg/l of free cyanide. In the latter case, the criteria approach those concentrations that will be protective of aquatic organisms, that is, approximately 5 µg/l (Virginia). New York is an exception and has a maximum allowable concentration of 400 µg/l of cyanide (unspecified form) as a groundwater effluent standard. Many

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states have developed groundwater quality guidelines for cyanide in the context of contaminated site remediation, however, as discussed in subsequent sections on RCRA and CERCLA.

18.1.3 FEDERAL WASTEWATER DISCHARGE STANDARDS The ambient water quality criteria provide guidance to states in adopting water quality standards, which ultimately provide a basis for controlling discharges or releases of pollutants into the waterways of the nation. The limits associated with these discharges or releases are set in the National Pollution Discharge Elimination System (NPDES) permits established under the authority of the Clean Water Act. Ambient water quality criteria that are derived to address site-specific situations are not included as part of this Federal regulation.

18.1.3.1 Effluent Guidelines and Standards Cyanide has been regulated in industrial wastewater discharges for many years. Under the direction of the Clean Water Act, the USEPA has developed effluent guidelines and standards for a large number of specific industries. These guidelines and standards are technology-based, that is, they are based on a projection of the effluent quality that will be produced by applying the best available treatment (BAT) technology to the typical wastewater that is generated by the specified industrial category. (Another category of treatment technologies is the best practical treatment or BPT. BPT differs from BAT in that the former gives some consideration to the cost of treatment.) These discharge standards have been developed for a total of 56 industry categories and are presented in the U.S. Code of Federal Regulations (40 CFR, Subchapter N, Parts 400 to 471). The discharge of some chemical form of cyanide is regulated in 13 of these industrial categories, listed in Table 18.6. Treatment technology evaluations for the industrial categories listed in Table 18.6 are implemented as part of the NPDES permit system, which requires a permit for all discharges to the surface waters of the nation. Permitted discharges must comply with the discharge limits that are prescribed for the appropriate industrial categories. State environmental agencies usually administer the NPDES program for the USEPA. In issuing a discharge permit, a state has the authority to stipulate either BAT limits or more stringent water quality limits, depending on the classification of the receiving water body. There is another set of discharge standards for these industrial categories which applies when the treated effluent is discharged to a publicly owned treatment facility, or POTW, prior to the discharge to a surface water body. These standards are known as pretreatment standards and were developed taking into consideration that some degree of treatment of the regulated contaminants would occur in the POTW. Two sets of standards exist for discharges to POTWs: (1) Pretreatment standards for new sources (PSNS) and (2) Pretreatment standards for existing sources (PSES). The differences in these standards reflect the assumption that new sources of wastewater are expected to generate reduced loads of contaminants as a result of improved or more efficient upstream process operations. Generally, these standards are essentially the same as BAT limits. For discharges to POTWs, pretreatment discharge limits may be based on PSES or PSNS limits or more stringent water quality limits, as dictated by NPDES requirements for the POTW. The Pollution Prevention Act, passed in 1990, is aimed at helping industry reduce or prevent pollution at the source, with one benefit being improved compliance with wastewater effluent guidelines and limits. USEPA was directed to provide technical assistance to businesses and to promote source reduction with industry. In response, USEPA developed initiatives with many different industries. In the context of reducing the volume and environmental impact of industrial wastewater discharges, the Agency initiated collaborations, for example, with the electroplating and metals manufacturing industries. Efforts of these industries in source reduction have yielded progress. In the years ahead there will be increasing focus on modification of manufacturing processes as part of wastewater

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TABLE 18.6 USEPA Industrial Wastewater Discharge Categories that Include Effluent Standards for Cyanide Electroplating Organic chemicals, plastics, and synthetic fibers Inorganic chemicals manufacturing Iron and steel manufacturing Nonferrous metals manufacturing Steam electric power generation Ferroalloy manufacturing Pharmaceutical manufacturing Photographic processing point sources Battery manufacturing Coil coating point sources Aluminum forming Nonferrous metals forming and metal powders point sources Note: Information from 40 CFR Chapter N Effluent Guidelines and Standards 400–471.

management strategies. The Pollution Prevention Act, and the assistance it makes available to companies through the USEPA, provides the framework for new directions in effluent limit compliance approaches. 18.1.3.2 Specific Characteristics of Guidelines/Standards 18.1.3.2.1 Specified cyanide analytical methods Part 136 of Title 40 (Guidelines Establishing Test Procedures for the Analysis of Pollutants) of the Code of Federal Regulations describes the analytical methods that should be used to determine compliance with the effluent guidelines and standards. Specifically, Table IB of Part 136 presents a list of approved inorganic test procedures. The test procedures that are specified for cyanide are: 1. Total cyanide: Manual distillation (Standard Methods 4500 CN C [5] and ASTM D2036-98(A) [6]) followed by titrimetric (Standard Methods 4500 CN D [5]) or spectrophotometric analysis, manual (USEPA Method 335.2 [7]; Standard Methods 4500 CN E [5]; ASTM 2036-98(A) [6], and USGS Method I-3300-85 [8]) or automated analysis (USEPA Method 335.3 [9] and USGS Method I-4302-85 [10]) of the distillation offgas absorber liquid; 2. Cyanide amenable to chlorination: Manual distillation with and without chlorination (USEPA Method 335.1 [11]; Standard Methods 4500 CN G [5]; and ASTM D2036-98(B) [12]) followed by titrimetric or spectrophotometric (manual or automated) analysis of the distillation offgas absorber liquid; and 3. Available cyanide: Flow injection and ligand exchange, followed by amperometry (USEPA Method OIA-1677 [13]). Based on these prescribed analytical methods, it is evident that the guidelines and standards are regulating industrial discharges on some combination of total, amenable, and available cyanide. Table 18.7 identifies those methods that have been prescribed for each of the 13 industrial subcategories that have specific discharge limits for cyanide. The specific discharge limits for each of the

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various categories are also presented in this table. Based upon a review of Table 18.7, it is clear that total cyanide is the predominant form of cyanide that is used for determining compliance, having been identified explicitly in 11 of the 13 industrial subcategories. For the other two industrial subcategories, no form of cyanide is specified. In many cases, amenable cyanide is also specified; however, it is used in tandem with the use of total cyanide. 18.1.3.2.2 Range of cyanide discharge limits The discharge limits for the 13 industrial categories that are presented in Table 18.7 are based on one of two criteria: (1) concentration-based criteria, that is, the cyanide concentration that is achievable in the effluent discharge after the application of a specific treatment, or (2) normalized mass criteria, that is, the mass of cyanide in the effluent discharge normalized per pound of product or related by-product. These different discharge criteria are also presented in terms of maximum values averaged over different time periods, typically one day, 4 days, or 30 days or monthly. Examples of the type and range of the concentration criteria shown in Table 18.7 are provided below: 1-day maximum: • Total cyanide: 1200 µg/l (organic chemicals, plastics, and synthetic fibers) — 33,500 µg/l (pharmaceutical manufacturing). • Amenable cyanide: 860 µg/l (metal finishing) — 5000 µg/l (electroplating national pretreatment standards for existing sources [<38, 000 l/day]). 4-day maximum averages: • Total cyanide: 1000 µg/l (electroplating national pretreatment standards for existing sources [>38, 000 l/day]). • Amenable cyanide: 2700 µg/l (electroplating national pretreatment standards for existing sources [<38, 000 l/day]). 30-day (monthly) maximum averages: • Total cyanide: 420 µg/l (organic chemicals, plastics, and synthetic fibers) — 9400 µg/l (pharmaceutical manufacturing). • Amenable cyanide: 320 µg/l (metal finishing). The normalized mass criteria in Table 18.7 cannot be easily summarized because of the different bases that were used for the normalization. For example, in the steam electric power generating subcategory, the allowable cyanide discharge is normalized per megawatt hour; in the precious metals subcategory, the discharge is normalized per troy ounce; while in several of the other categories, it is per million pounds of material production.

18.1.4 GROUNDWATER (RCRA AND CERCLA) Two primary pieces of U.S. legislation that govern the management of residuals containing cyanide, and groundwater contacted by these residuals, are the Resource Conservation and Recovery Act (RCRA) and the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Congress enacted RCRA in 1976 (RCRA Public Law 94-580) and CERCLA in 1980 (later amended and reauthorized in 1986, Public Law 99-499). RCRA established a system for managing hazardous wastes from the point of origin to the final disposal, that is, cradle to grave, while CERCLA addresses legacy sites where hazardous substances have been released to the environment. 18.1.4.1 Groundwater (RCRA) Under RCRA, wastes may be classified as hazardous wastes or solid (nonhazardous wastes). A solid waste under RCRA is hazardous if it is not excluded from the hazardous waste regulations and (1) it

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Total cyanide

Organic chemicals, plastics, and synthetic fibers (with and without end-of-pipe biological treatment) Inorganic chemicals manufacturing (1) Hydrogen peroxide production (electrolytic process)

(2) Hydrogen cyanide production

Total cyanide

(2) ≥38,000 l/day

Total cyanidec

Amenable cyanide

Amenable cyanide

Amenable cyanide

Cyanide form

Electroplating (1) <38,000 l/day

Industrial subcategory

0.10 lbs/1000 lbs of product (1-day max) 0.021 lbs/1000 lbs of product (30-day average) 0.65 lbs/1000 lbs of product (1-day max) 0.23 lbs/1000 lbs of product (30-day average)

NSPS

11,000 µg/l (1-day max) 4000 µg/l (30-day average)

5.0 mg/l (1-day max) 2.7 mg/l (4-day average max) 1.9 mg/l (1-day max) 1.0 mg/l (4-day average max) 1200 µg/l (1-day maximum) 420 µg/l (max monthly average)

PSES

TABLE 18.7 Technology Performance Standards for Selected Industrial Categoriesa

0.65 lbs/1000 lbs of product (1-day max) 0.23 lbs/1000 lbs of product (30-day average)

1700 µg/l (1-day max); 360 µg/l (30-day average)

1200 µg/l (1-day maximum) 420 µg/l (max monthly average)

Standardb PSNS

0.0004 lbs/1000 lbs of product (1-day max) 0.0002 lbs/1000 lbs of product (30-day average) 0.10 lbs/1000 lbs of product (1-day max) 0.021 lbs/1000 lbs of product (30-day average) 0.65 lbs/1000 lbs of product (1-day max) 0.23 lbs/1000 lbs of product (30-day average)

BPT

0.10 lbs/1000 lbs of product (1-day max) 0.021 lbs/1000 lbs of product (30-day average)

1200 µg/l (1-day maximum) 420 µg/l (max monthly average)

BAT

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Total cyanide

Total cyanide

(b) Continuous

0.00102 lbs/1000 lbs of product (1-day max) 0.000339 lbs/1000 lbs of product (30-day average) 0.00569 lbs/1000 lbs of product (1-day max) 0.00190 lbs/1000 lbs of product (30-day average)

0.000584 lbs/1000 lbs of product (1-day average) 0.000292 lbs/1000 lbs of product (30-day average)

Total cyanide

Total cyanide

0.00297 lbs/1000 lbs of product (1-day max) 0.00208 lbs/1000 lbs of product (max monthly average) 0.00100 lbs/1000 lbs of product (1-day max) 0.000501 lbs/1000 lbs of product (max monthly average)

NSPS

Total cyanide

Cyanide form

Salt bath descaling (Reducing) (a) Batch

(2) Sintering (NSPS, PSES, and PSNS apply only when sintering wastewater is cotreated with iron making wastewater) (3) Ironmaking (iron blast furnace)

Iron/steel manufacturingd (1) By-product coke making (iron and steel)

Industrial subcategory

TABLE 18.7 Continued

0.00102 lbs/1000 lbs of product (1-day max) 0.000339 lbs/1000 lbs of product (30-day average) 0.00569 lbs/1000 lbs of product (1-day max) 0.00190 lbs/1000 lbs of product (30-day average)

0.00175 lbs/1000 lbs of product (1-day average) 0.000876 lbs/1000 lbs of product (30-day average)

0.00724 lbs/1000 lbs of product (1-day max) 0.00506 lbs/1000 lbs of product (max monthly average) 0.00300 lbs/1000 lbs of product (1-day max) 0.00150 lbs/1000 lbs of product (max monthly average)

PSES

0.00102 lbs/1000 lbs of product (1-day max) 0.000339 lbs/1000 lbs of product (30-day average) 0.00569 lbs/1000 lbs of product (1-day max) 0.00190 lbs/1000 lbs of product (30-day average)

0.000584 lbs/1000 lbs of product (1-day average) 0.000292 lbs/1000 lbs of product (30-day average)

0.00297 lbs/1000 lbs of product (1-day max) 0.00208 lbs/1000 lbs of product (max monthly average) 0.00100 lbs/1000 lbs of product (1-day max) 0.000501 lbs/1000 lbs of product (max monthly average)

Standard,b PSNS

0.00102 lbs/1000 lbs of product (1-day max) 0.000339 lbs/1000 lbs of product (30-day average) 0.00569 lbs/1000 lbs of product (1-day max) 0.00190 lbs/1000 lbs of product (30-day average)

0.0234 lbs/1000 lbs of product (1-day average) 0.00782 lbs/1000 lbs of product (30-day average)

0.0657 lbs/1000 lbs of product (1-day max) 0.0219 lbs/1000 lbs of product (max monthly average) Cyanide limits not specified

BPT

0.00102 lbs/1000 lbs of product (1-day max) 0.000339 lbs/1000 lbs of product (30-day average) 0.00569 lbs/1000 lbs of product (1-day max) 0.00190 lbs/1000 lbs of product (30-day average)

0.00175 lbs/1000 lbs of product (1-day average) 0.000876 lbs/1000 lbs of product (30-day average)

0.00297 lbs/1000 lbs of product (1-day max) 0.00208 lbs/1000 lbs of product (max monthly average) 0.00300 lbs/1000 lbs of product (1-day max) 0.00150 lbs/1000 lbs of product (max monthly average)

BAT

364 Cyanide in Water and Soil

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Total cyanide

Total cyanide

(3) Secondary precious metals

Total cyanide

Total cyanide

Total cyanide

(2) Primary beryllium

(b) Cathode reprocessing operated with wet potline scrubbing (c) Potline wet air pollution control operated with cathode reprocessing

Nonferrous metals manufacturing (1) Primary aluminum smelting (a) Cathode reprocessing with dry potline scrubbing

0.00 to 449.2 lbs/million lbs of beryllium carbonate produced (1-day max) 0.0 to 179.7 lbs/million lbs of beryllium carbonate produced (max monthly average) 0.00 to 10.0 mg/troy oz. (1-day max) 0.00 to 4.0 mg/troy oz. (max monthly average)

157.6 lbs/million lbs of cryolite recovered (1-day max) 70.06 lbs/million lbs of cryolite recovered (max monthly average)

0.00 to 10.0 mg/troy oz. (1-day max) 0.00 to 4.0 mg/ troy oz. (max monthly average)

157.6 lbs/million lbs of cryolite recovered (1-day max) 70.06 lbs/million lbs of cryolite recovered (max monthly average)

0.00 to 449.2 lbs/million lbs of beryllium carbonate produced (1-day max) 0.0 to 179.7 lbs/million lbs of beryllium carbonate produced (max monthly average) 0.00 to 10.0 mg/troy oz. (1-day max) 0.00 to 4.0mg/troy oz. (max monthly average)

0.00 to 651.3 lbs/million lbs of beryllium carbonate produced (1-day max) 0.0 to 269.5 lbs/million lbs of beryllium carbonate produced (max monthly average) 0.00 to 20.82 mg/ troy oz. (1-day max) 0.00 to 8.616 mg/ troy oz. (max monthly average)

3.771 lbs/million lbs of aluminum produced (1-day max) 1.676 lbs/million lbs of aluminum produced (30-day average) 0.00 to 449.2 lbs/million lbs of beryllium carbonate produced (1-day max) 0.0 to 179.7 lbs/million lbs of beryllium carbonate produced (max monthly average) 0.00 to 10.0 mg/troy oz. (1-day max) 0.00 to 4.0 mg/ troy oz. (max monthly average)

157.6 lbs/million lbs of cryolite recovered (1-day max) 70.06 lbs/million lbs of cryolite recovered (max monthly average) 0.00 (1-day max and max monthly average)

Regulation of Cyanide in Water and Soil 365

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Total cyanide

(3) Metal finishing

Amenable cyanide

Total cyanide

(2) Covered calcium carbide furnace

0.0005 kg/Mwh (1-day average) 0.0003 kg/Mwh (30-day average)

Nondetect in chemicals Added for cooling tower maintenance

Total cyanide

Steam electric power generation

Total cyanide

0.00 to 8.694 mg/kg (1-day max) 0.00 to 3.478 mg/kg (max monthly average)

Total cyanide

(5) Primary zirconium and hafnium

Ferroalloy manufacturing (1) Covered electric furnace

0.007 to 23.0 mg/kg (1-day max) 0.003 to 9.2 mg/kg (max monthly average)

NSPS

Total cyanide

Cyanide form

(4) Secondary tin

Industrial subcategory

TABLE 18.7 Continued

1.20 mg/l (1-day max) 0.65 mg/l (30-day average) 0.86 mg/l (1-day max) 0.32 mg/l (30-day average)

Nondetect in chemicals Added for cooling tower maintenance

0.007 to 23.0 mg/kg (1-day max) 0.003 to 9.2 mg/kg (max monthly average)

PSES

1.20 mg/l (1-day max) 0.65 mg/l (30-day average) 0.86 mg/l (1-day max) 0.32 mg/l (30-day average)

Nondetect in chemicals Added for cooling tower maintenance

0.00 to 8.694 mg/kg (1-day max) 0.00 to 3.478 mg/kg (max monthly average)

0.007 to 23.0 mg/kg (1-day max) 0.003 to 9.2 mg/kg (max monthly average)

Standardb PSNS

0.004 kg/Mwh (1 day max) 0.002 kg/Mwh (30-day average) 0.0056 lbs/million lbs (1-day max) 0.0028 lbs/million lbs (30-day average) 1.20 mg/l (1-day max) 0.65 mg/l (30-day average) 0.86 mg/l (1-day max) 0.32 mg/l (30-day average)

0.00 to 12.610 mg/kg (1-day max) 0.00 to 5.216 mg/kg (max monthly average)

0.010 to 33.35 mg/kg (1-day max) 0.004 to 13.8 mg/kg (max monthly average)

BPT

0.0005 kg/Mwh (1-day average) 0.0003 kg/Mwh (30-day average) 0.0056 lbs/million lbs (1-day max) 0.0028 lbs/million lbs (30-day average) 1.20 mg/l (1-day max) 0.65 mg/l (30-day average) 0.86 mg/l (1-day max) 0.32 mg/l (30-day average)

0.007 to 23.0 mg/kg (1-day max) 0.003 to 9.2 mg/kg (max monthly average) 0.00 to 8.694 mg/kg (1-day max) 0.00 to 3.478 mg/kg (max monthly average) Nondetect in chemicals Added for cooling tower maintenance

BAT

366 Cyanide in Water and Soil

Total cyanide

Not specified

(3) Aluminum basis material

Aluminum forming (1) Rolling with neat oil

Not specified

Not specified

(2) Galvanized basis material

Coil coating (1) Steel basis material

Total cyanide

Not specified

Photographic processingd

Battery manufacturing: zinc subcategory

Total cyanide

Pharmaceutical manufacturing

0.00039 to 0.41 lbs/ million lbs of aluminum rolled (1-day average) 0.00016 to 0.17 lbs/ million lbs of aluminum rolled (max monthly average)

0.063 mg/m2 of area processed (1-day max) 0.025 mg/m2 of area processed (max monthly average) 0.07 mg/m2 of area processed (1-day max) 0.028 mg/m2 of area processed (max monthly average) 0.095 mg/m2 of area processed (1-day max) 0.038 mg/m2 of area processed (max monthly average)

0.039 mg/kg (1-day max) 0.016 mg/kg (30-day average)

33.5 mg/l (1-day max) 9.4 mg/l (30-day average)

0.00057 to 0.59 lbs/ million lbs of aluminum rolled (1-day average) 0.00024 to 0.25 lbs/ million lbs of aluminum rolled (max monthly average)

0.34 mg/m2 of area processed (1-day max) 0.14 mg/m2 of area processed (max monthly average) 0.26 mg/m2 of area processed (1-day max) 0.11 mg/m2 of area processed (max monthly average) 0.29 mg/m2 of area processed (1-day max) 0.12 mg/m2 of area processed (max monthly average)

0.38 mg/kg (1-day max) 0.16 mg/kg (30-day average)

33.5 mg/l (1-day max) 9.4 mg/l (30-day average)

0.00039 to 0.41 lbs/ million lbs of aluminum rolled (1-day average) 0.00016 to 0.17 lbs/ million lbs of aluminum rolled (max monthly average)

0.063 mg/m2 of area processed (1-day max) 0.025 mg/m2 of area processed (max monthly average) 0.07 mg/m2 of area processed (1-day max) 0.028mg/m2 of area processed (max monthly average) 0.095 mg/m2 of area processed (1-day max) 0.038 mg/m2 of area processed (max monthly average)

0.039 mg/kg (1-day max) 0.016 mg/kg (30-day average)

33.5 mg/l (1-day max) 9.4 mg/l (30-day average)

0.00057 to 4.61 lbs/ million lbs of aluminum rolled (1-day average) 0.00024 to 1.91 lbs/ million lbs of aluminum rolled (max monthly average)

0.80 mg/m2 of area processed (1-day max) 0.33mg/m2 of area processed (max monthly average) 0.76 mg/m2 of area processed (1-day max) 0.32 mg/m2 of area processed (max monthly average) 0.98mg/m2 of area processed (1-day max) 0.41mg/m2 of area processed (max monthly average)

33.5 mg/l (1-day max) 9.4 mg/l (30-day average) 0.038lb/1000 ft2 of product (91-day max) 0.019lb/1000 ft2 of product (max monthly average) 2.54 mg/kg (1-day max) 1.05 mg/kg (30-day average)

0.00057 to 4.04 lbs/ million lbs of aluminum rolled (1-day average) 0.00024 to 1.67 lbs/ million lbs of aluminum rolled (max monthly average)

0.34mg/m2 of area processed (1-day max) 0.14mg/m2 of area processed (max monthly average) 0.26 mg/m2 of area processed (1-day max) 0.11 mg/m2 of area processed (max monthly average) 0.29mg/m2 of area processed (1-day max) 0.12mg/m2 of area processed (max monthly average)

0.38 mg/kg (1-day max) 0.16 mg/kg (30-day average)

33.5 mg/l (1-day max) 9.4 mg/l (30-day average)

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0.026 to 0.41 lbs/million lbs of aluminum rolled (1-day average) 0.011 to 0.16 lbs/million lbs of aluminum rolled (max monthly average)

0.036 to 0.41 lbs/million lbs of aluminum extruded (1-day max) 0.024 to 0.17 lbs/million lbs of aluminum extruded (max monthly average) 0.010 to 0.41 lbs/million lbs of aluminum forged (1-day average) 0.004 to 0.163 lbs/million lbs of aluminum forged 0.0004 to 0.408 lbs/ million lbs of aluminum drawn (1-day average) 0.0002 to 0.163 lbs/ million lbs of aluminum drawn (max monthly average)

Total cyanide

Total cyanide

Total cyanide

(3) Extrusion

(4) Forging

(5) Drawing with neat oil

NSPS

Total cyanide

Cyanide form

(2) Rolling with emulsions

Industrial subcategory

TABLE 18.7 Continued

0.0006 to 0.591 lbs/ million lbs of aluminum drawn (1-day average) 0.0003 to 0.245 lbs/ million lbs of aluminum drawn (max monthly average)

0.052 to 1.2 lbs/million lbs of aluminum extruded (1-day max) 0.022 to 0.5 lbs/million lbs of aluminum extruded (max monthly average) 0.015 to 1.2 lbs/million lbs of aluminum forged (1-day average) 0.006 to 0.5 lbs/million lbs of aluminum forged

0.038 to 0.59 lbs/million lbs of aluminum rolled (1-day average) 0.016 to 0.25 lbs/million lbs of aluminum rolled (max monthly average)

PSES

0.036 to 0.41 lbs/million lbs of aluminum extruded (1-day max) 0.015 to 0.17 lbs/million lbs of aluminum extruded (max monthly average) 0.010 to 0.41 lbs/million lbs of aluminum forged (1-day average) 0.004 to 0.163 lbs/million lbs of aluminum forged 0.0004 to 0.408/ million lbs of aluminum drawn (1-day average) 0.0002 to 0.163 lbs/ million lbs of aluminum drawn (max monthly average)

0.026 to 0.41 lbs/million lbs of aluminum rolled (1-day average) 0.011 to 0.16 lbs/million lbs of aluminum rolled (max monthly average)

Standardb PSNS

0.00057 to 4.61 lbs/ million lbs of aluminum drawn (1-day average) 0.00024 to 1.91 lbs/ million lbs of aluminum drawn (max monthly average)

0.052 to 4.61 lbs/million lbs of aluminum extruded (1-day max) 0.022 to 1.91 lbs/million lbs of aluminum extruded (max monthly average)

0.038 to 4.61 lbs/million lbs of aluminum rolled (1-day average) 0.016 to 1.91 lbs/million lbs of aluminum rolled (max monthly average)

BPT

0.0006 to 0.591 lbs/ million lbs of aluminum drawn (1-day average) 0.0002 to 0.245 lbs/ million lbs of aluminum drawn (max monthly average)

0.038 to 0.59 lbs/ million lbs of aluminum rolled (1-day average) 0.016 to 0.25 lbs/million lbs of aluminum rolled (max monthly average) 0.052 to 1.2 lbs/million lbs of aluminum extruded (1-day max) 0.022 to 0.5 lbs/million lbs of aluminum extruded (max monthly average)

BAT

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Not specified

Not specified

(3) Zinc forming

Not specified

Total cyanide

(2) Titanium forming

Nonferrous metals forming and metal powders point sources (1) Precious metal forming

(6) Drawing with emulsions of soap

0.0009 to 1.94 lbs/ million lbs of metal formed (1-day average) 0.0004 to 0.802 lbs/ million lbs of metal formed (max monthly average) 0.010 to 0.84 lbs/million lbs of titanium formed (1-day average) 0.005 to 0.351 lbs/ million lbs of titanium formed (max monthly average) 0.0003 to 0.338 lbs/million lbs of zinc formed (1-day average) 0.0001 to 0.135 lbs/million lbs of zinc formed (max monthly average)

0.0004 to 0.408 lbs/ million lbs of aluminum drawn (1-day average) 0.0002 to 0.16 lbs/ million lbs of aluminum drawn (max monthly average)

0.0009 to 1.94 lbs/ million lbs of metal formed (1-day average) 0.0004 to 0.802 lbs/ million lbs of metal formed (max monthly average) 0.010 to 0.84 lbs/million lbs of titanium formed (1-day average) 0.005 to 0.351 lbs/ million lbs of titanium formed (max monthly average)

0.0006 to 0.591 lbs/ million lbs of aluminum drawn (1-day average) 0.0003 to 0.25 lbs/ million lbs of aluminum drawn (max monthly average)

0.0009 to 1.94 lbs/ million lbs of metal formed (1-day average) 0.0004 to 0.802 lbs/ million lbs of metal formed (max monthly average) 0.010 to 0.84 lbs/million lbs of titanium formed (1-day average) 0.005 to 0.351 lbs/ million lbs of titanium formed (max monthly average) 0.0003 to 0.338 lbs/ million lbs of zinc formed (1-day average) 0.0001 to 0.135 lbs/ million lbs of zinc formed (max monthly average)

0.0004 to 0.408 lbs/ million lbs of aluminum drawn (1-day average) 0.0002 to 0.16 lbs/ million lbs of aluminum drawn (max monthly average)

0.0004 to 1.04 lbs/ million lbs of zinc formed (1-day average) 0.0002 to 0.430 lbs/ million lbs of zinc formed (max monthly average)

0.0009 to 3.51 lbs/ million lbs of metal formed (1-day average) 0.0004 to 1.45 lbs/ million lbs of metal formed (max monthly average) 0.010 to 8.47 lbs/million lbs of titanium formed (1-day average) 0.004 to 3.51 lbs/million lbs of titanium formed (max monthly average)

0.0006 to 4.61 lbs/ million lbs of aluminum drawn (1-day average) 0.0003 to 1.91 lbs/ million lbs of aluminum drawn (max monthly average)

0.0009 to 1.94 lbs/ million lbs of metal formed (1-day average) 0.0004 to 0.802 lbs/ million lbs of metal formed (max monthly average) 0.010 to 0.84 lbs/million lbs of titanium formed (1-day average) 0.005 to 0.351 lbs/ million lbs of titanium formed (max monthly average) 0.0003 to 0.338 lbs/million lbs of zinc formed (1-day average) 0.0001 to 0.135 lbs/ million lbs of zinc formed (max monthly average)

0.0006 to 0.591 lbs/ million lbs of aluminum drawn (1-day average) 0.0003 to 0.25 lbs/ million lbs of aluminum drawn (max monthly average)

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Not specified

(5) Metals powder 0.004 to 2.29 lbs/million lbs of powder atomized (1-day average) 0.002 to 0.948 lbs/ million lbs of powder atomized (max monthly average)

0.005 to 9.11 lbs/million lbs of metal formed (1-day average) 0.002 to 3.77 lbs/million lbs of zirconium-hafnium formed (max monthly average)

NSPS

0.004 to 2.55 lbs/million lbs of powder atomized (1-day average) 0.002 to 1.06 lbs/million lbs of powder atomized (max monthly average)

0.005 to 9.11 lbs/million lbs of metal formed (1-day average) 0.002 to 3.77 lbs/million lbs of zirconium-hafnium formed (max monthly average)

PSES

0.004 to 2.29 lbs/million lbs of powder atomized (1-day average) 0.002 to 0.948 lbs/million lbs of powder atomized (max monthly average)

0.005 to 9.11 lbs/million lbs of metal formed (1-day average) 0.002 to 3.77 lbs/million lbs of zirconium-hafnium formed (max monthly average)

Standardb PSNS

0.004 to 2.55 lbs/million lbs of powder atomized (1-day average) 0.002 to 1.06 lbs/million lbs of powder atomized (max monthly average)

0.005 to 9.11 lbs/million lbs of metal formed (1-day average) 0.002 to 3.77 lbs/million lbs of zirconium-hafnium formed (max monthly average)

BPT

0.005 to 9.11 lbs/million lbs of metal formed (1-day average) 0.002 to 3.77 lbs/million lbs of zirconium-hafnium formed (max monthly average) 0.004 to 2.55 lbs/million lbs of powder atomized (1-day average) 0.002 to 1.06 lbs/million lbs of powder atomized (max monthly average)

BAT

technology. BAT = Best available treatment technology. c 40 CFR Subchapter N Part 415.422–415.426. d Information from USFDA, 40 CFR Part 420, Effluent limitations guidelines, pretreatment standards and new source performance standards for the iron and steel manufacturing point source category; Final rule, U.S. Environmental Protection Agency, Washington, DC, 2002.

b NSPS = New source performance standards. PSES = Pretreatment standards for existing sources. PSNS = Pretreatment standards for few sources. BPT = Best practical treatment

Registry, Atlanta, GA, 1997.

a Information from ASTDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease

Not specified

Cyanide form

(4) Zirconium–hafnium forming

Industrial subcategory

TABLE 18.7 Continued

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TABLE 18.8 Land Disposal Treatment Standards for Cyanide in Water Waste description under RCRA D003 F006, F019 F007, F008, F009 F010 F011, F012, P013, P021, P029, P030, P063, P074, P098, P099, P104, P106, P121 F037, K048, K049, K050, K051, K052 K005, K007 K011, K013, K014 K060 K086

Cyanide form

Standard (mg/l)

Amenable cyanide Total cyanide Total cyanide Amenable cyanide Total cyanide Amenable cyanide Total cyanide Amenable cyanide Total cyanide

0.86 Reserved 1.20 0.86 1.90 0.100 1.90 0.100 1.90

Amenable cyanide Total cyanide

0.100 0.028

Total cyanide Total cyanide Total cyanide Total cyanide

0.740 21.0 1.90 1.90

Source: Information from ATSDR, Toxicological profile for Cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

is a listed waste (i.e., listed in one of three lists [“F” List — Nonspecific source waste; “K” List — specific source list; or “P” or “U” Lists — Commercial chemical products] developed by USEPA and contained in the Code of Federal Regulations (CFR) at 40 CFR 261.31-33), or (2) exhibits one or more of four characteristics; ignitability, corrosivity, reactivity, and toxicity (40 CFR 261.21.24). If an environmental medium, such as groundwater, becomes contaminated (i.e., mixed) with a hazardous waste, in accordance with the USEPA “contained in” policy, it will be regulated as a hazardous waste until such time as the medium is treated to remove the contaminant. This policy results in the potential applicability of the Land Disposal Restrictions (LDR) to contaminated media during environmental remediation activities. The LDRs specify concentration levels for contaminants that must be achieved before the impacted material can be placed in a landfill, surface impoundment, waste pile, injection well, land treatment facility, salt dome formation, underground mine or cave, or concrete bunker or vault. (LDRs do not apply to wastes that are discharged to surface waters, where NPDES requirements apply, or to Publicly Owned Treatment Works, where pretreatment requirements apply.) There are several land disposal treatment standards that would be applicable to impacted groundwater should it become mixed with specific listed or characteristic hazardous wastes that contain cyanide. These concentration limits and the form of cyanide to which they apply are summarized in Table 18.8. Many states have been active in developing guidelines and standards for groundwater remediation under RCRA and CERCLA. A wide range of approaches has been used in establishing these guidelines and standards. As cyanide is a frequently occurring subsurface contaminant [14], it is frequently included in state lists of guidelines and standards. These lists are available via the Internet; an excellent compendium is available at cleanuplevels.com [15]. In Pennsylvania, for example,

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groundwater “medium specific concentrations” (MSCs) have been developed for free cyanide. The free cyanide MSC is 200 µg/l for used aquifers with total dissolved solids (TDS) less than 2,500 mg/l; 20,000 µg/l for used aquifers with TDS greater than 2,500 mg/l; and 200,000 µg/l for nonuse aquifers. Wisconsin has developed state public health groundwater quality standards for use in remediation planning. The “enforcement standard” for cyanide (presumably free cyanide, but form not specified) is 200 µg/l, while the “preventive action limit” is 40 µg/l. As is evident from these two examples, many of the guidelines or standards are tied to the USEPA drinking water MCL of 200 µg/l. 18.1.4.2 Groundwater (CERCLA) Section 121(d) of the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA), as amended by the 1986 Superfund Amendments and Reauthorization Act (SARA), requires that onsite remedial actions must attain or waive Federal or more stringent State applicable or relevant and appropriate requirements (ARARs) upon completion of the remedial action. There are three types of ARARs: (1) chemical specific, for example, total cyanide concentration in soil <50 mg/kg; (2) action specific, for example, if an onsite landfill is proposed, it should meet contemporary landfill standards; and (3) location specific, for example, prohibition of land disposal in a flood plain. Where ARARs do not exist for important chemicals at a site, or where the USEPA determines an environmental regulation is not appropriate to determine clean-up levels, such levels may be set through the use of quantitative risk assessment. More specifically for impacted groundwater at a CERCLA site, maximum contaminant levels (MCLs) and maximum contaminant level goals (MCLGs) under the Safe Drinking Water Act and water quality criteria under the Clean Water Act represent potential chemical-specific ARARs. At the McAdoo Associates Superfund Site in Pennsylvania, for example, the USEPA acute-exposure ambient water quality criterion for free cyanide of 22 µg/l was adopted as the groundwater quality to be achieved at a downgradient point of exposure [16]. Location-specific ARARs might include limits on activities in wetlands as prescribed under Sections 401 and 404 of the Clean Water Act. Lastly, action-specific ARARs are usually technology or activity-based requirements or limitations on actions or conditions involving specific substances such as the application of existing BPT or BAT to cyanide-impacted groundwater. Stated differently, nearly all of the water-related regulatory requirements that are discussed within this chapter may be applicable, under some set of circumstances, to a groundwater on a CERCLA site that has been impacted by some form of cyanide. As noted in the previous discussion on RCRA, states have been very active in developing guidelines and standards for groundwater remediation. As cyanide is a frequently occurring subsurface contaminant [14], it is frequently included in state lists of guidelines and standards. These lists are available at cleanuplevels.com [15].

18.2 U.S. REGULATIONS, GUIDELINES, AND CRITERIA FOR SOIL, SEDIMENT, AND PROCESS RESIDUALS 18.2.1 SOIL AND PROCESS RESIDUALS As previously mentioned, wastes may be classified as hazardous wastes or solid (nonhazardous) wastes. A solid waste under RCRA is hazardous if it is not excluded from the hazardous waste regulations and (1) it is a listed waste (i.e., listed in one of three lists [“F” List, “K” List, or the “P” or “U” Lists]), or (2) exhibits one or more of four characteristics, ignitability, corrosivity, reactivity, and toxicity. Similar to groundwater, if soil becomes contaminated (i.e., mixed) with a hazardous waste, in accordance with the USEPA “contained in” policy, it will be regulated as a hazardous waste until

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such time as it is treated to remove the contaminant. This policy results in the potential applicability of the Land Disposal Restrictions (LDR) to contaminated media during environmental remediation activities. The LDRs specify concentration levels for contaminants that must be achieved before the impacted soil can be placed in a landfill, waste pile, land treatment facility, concrete bunker, or vault. In the context of RCRA and CERCLA, the states have been very active in developing guidelines and standards for soil remediation. A wide range of approaches has been used in establishing these guidelines and standards. As cyanide is a frequently occurring subsurface contaminant [14], it is frequently included in state lists of guidelines and standards. These lists are available at cleanuplevels.com [15]. As an example, Pennsylvania has established “medium specific concentrations” (MSCs) for cyanide in soil for use in direct contact risk evaluations, and also for soil-to-groundwater evaluations. The direct contact MSCs for free cyanide are 4,400 mg/kg for surface soil at residential sites (0 to 15 ft depth); 56,000 mg/kg for surface soil at nonresidential sites (0 to 2 ft depth); and 190,000 mg/kg for subsurface soil at nonresidential sites (2 to 15 ft depth). The free cyanide MSCs for the soil-to-groundwater scenario are 200 mg/kg for both residential and nonresidential sites when impacts on a used aquifer with TDS <2, 500 mg/l are relevant; 2,000 mg/kg for both residential and nonresidential sites when impacts on a used aquifer with TDS >2,500 mg/l are relevant; and 190,000 mg/kg for both residential and nonresidential sites when impacts on a nonuse aquifer are relevant. Other states have established their own kinds of classifications considering the unique nature and uses of their soils and groundwaters. 18.2.1.1 Listed Wastes There are several listed wastes that contain cyanide and for which land disposal treatment standards have been specified. The same land disposal restrictions apply to soil that has been mixed with these wastes. Table 18.9 provides a list of these cyanide-containing wastes and the concentrations of cyanide that must be achieved before they, or a mixture of the waste with soil, can be managed in one of the above-noted land disposal units. The data provided in Table 18.9 indicate that concentrations of amenable cyanide between 9 and 30 mg/kg can be placed in a land disposal facility while total cyanide concentrations in a land disposal facility are limited to between 1.2 and 590 mg/kg. 18.2.1.2 Characteristic Wastes Reactivity is one of the four “characteristics” defined in RCRA that must be satisfied for a solid waste to be considered a “characteristic” hazardous waste. Until recently, one of the categories of wastes defined in the regulations in 40CFR 261.23 as a reactive waste had the following properties: cyanide-bearing waste, which generate toxic fumes when exposed to mild acidic or basic conditions. However, in 2004 this regulation was revised by withdrawing the category of reactive cyanide. This withdrawal of the reactive cyanide was prompted by the fact that the specified test conditions, that is, exposure to a strong acid, would rarely, if ever, yield toxic cyanide fumes because of the extremely low solubility of the common solid forms of cyanide, such as ferric ferrocyanide (see Chapters 2 and 5), in water at pH values in the range 2.0 to 5.0.

18.2.2 SEDIMENT Sediments in freshwater, estuarine, and coastal systems can become contaminated as a result of interactions with the overlying water and from deposition of material already contaminated. The presence of chemical contaminants in sediments is a concern with respect to potential effects on benthic organisms living in sediments, and on the broader community of aquatic organisms, especially those that feed on benthic organisms and their food chain relatives. Since contaminated sediments serve as a reservoir of chemical mass that can be released to the water column over time, the entire

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TABLE 18.9 Land Disposal Treatment Standards for Cyanide in Soil/Wastes Waste description under RCRA D003 F006, F007, F008, F009, F019 F010 F011, F012, P013, P021, P029, P030, P063, P074, P098, P099, P104, P106, P121 F037, K048, K049, K050, K051, K052 K005, K007 K011, K013, K014 K060 K086

Cyanide form

Standard (mg/kg)

Amenable cyanide Total cyanide Amenable cyanide Total cyanide Total cyanide Amenable cyanide Total cyanide

30 590 30 590 1.5 Not applicable 110

Amenable cyanide Total cyanide

9.1 1.8

Total cyanide Total cyanide Total cyanide Total cyanide

Reserved 57 1.2 1.5

Source: Information from ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997.

aquatic community of plants and animals is potentially impacted by the presence of contaminants in sediments. Chemical quality standards or guidelines for sediments are most relevant in three primary contexts in the United States: (1) management of sediment from navigational dredging; (2) cleanup actions under CERCLA in which sediments have been identified as a contaminated site or part of a contaminated site; and (3) development of total maximum daily load (TMDL) plans for protection of aquatic systems under the Clean Water Act. For dredging operations and contaminated site remediation, quality objectives are established on a site-specific basis, using state or U.S. quality guidelines or site-specific risk assessment. These assessments make use of existing water quality and sediment quality criteria established for the protection of aquatic life. Thus, sediment management or cleanup objectives are related to the same analyses performed to identify chemical levels that are protective of benthic and aquatic life. Efforts to plan for and develop U.S. sediment quality criteria have been underway since the 1970s. In addition to the scientific complexities involved, the development of sediment quality criteria has been administratively challenging in that more than ten Federal statutes provide authority to USEPA to address contaminated sediment [17]. Under Section 304(a) of the Clean Water Act, the USEPA is specifically charged with the development and implementation of sediment quality criteria. Implementation of this charge, however, requires coordination with regulations developed under other statutes that pertain to dredging, contaminated site remediation, and other areas. USEPA has developed a Contaminated Sediment Management Strategy to address the coordination issue and to establish the framework for development of sediment quality criteria and approaches to management of contaminated sediment [17]. The USEPA is developing sediment quality criteria for the protection of benthic organisms, to be applied in cases where the sediment total organic carbon exceeds 0.2% dry weight, the primary route of exposure is in direct contact with the sediment, and the sediments are continually submerged

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or there is information indicating that equilibrium has been established between the water and the sediments. Sediment quality criteria have been developed for several nonionic organic compounds, and are in development for additional organic compounds and selected metals. No sediment quality criteria have been established by the USEPA for any cyanide species. As part of the Contaminated Sediment Management Strategy, USEPA is conducting a National Sediment Inventory [18] for the purpose of assessing the quality of U.S. freshwater, estuarine, and coastal sediments, and for building the scientific foundation upon which national sediment quality criteria can be established. Sediment quality screening levels have been established for some contaminants by the USEPA, for the purpose of interpreting the NSI data [18]. These screening levels are for selected metals and persistent organic compounds. No screening levels have been established yet for any cyanide species. It is not the intention of the USEPA that sediment quality criteria or screening levels should be used as mandatory cleanup levels for remediation or decision levels for dredged material management [17]. Decisions about what constitutes desired or acceptable sediment quality in those situations should be based on analysis of potential effects for different levels and types of contaminants in particular situations. This is the governing principle of the U.S. Army Corps of Engineers in their management of dredged material [19]. A number of states, for example, Florida and Washington, have developed sediment quality criteria. It appears that no state sediment quality criteria have been established yet for any cyanide species. As is the case for the USEPA sediment quality screening levels, the state sediment quality criteria that have been established are focused primarily on some metals and persistent organic compounds. Some cyanide sediment quality guidelines have been developed and reported in the United States for specific applications. These are summarized in Table 18.10. For example, in the 1970s USEPA Region 5 developed some guidelines for cyanide concentrations in nonpolluted, moderately-polluted, and heavily-polluted sediments [20]. The sediment quality guidelines reported in Table 18.10, while not developed in the current framework for establishment of U.S. sediment quality criteria, suggest that concentrations below 0.1 mg/kg will be protective of benthic life in freshwater systems. For the guidelines summarized in Table 18.10, the particular form of cyanide was not specified in the original reports. A conservative approach would be to consider these values to refer to total cyanide, though the toxic form of interest will usually be free cyanide. Cyanide contamination in sediments is not often a driver in sediment management regulation and decision-making because dissolved free cyanide (HCN, CN− ), the most toxic form of cyanide, is highly soluble and biodegradable (see Chapters 5 and 6) and thus usually has a short half-life in sediments. Other dissolved forms of cyanide, most notably metal–cyanide complexes, can be retained longer in sediments but are less toxic. In addition, cyanide species have low bioaccumulation potential (Chapter 14; and [4]). Cyanide in sediments may become a concern when present in significant mass concentrations and total amounts, such as will be the case for sediments contaminated from a former waste disposal activity or spill. Development of a sediment quality goal for cyanide at a particular site, if desired, will require a site-specific risk assessment. If such an assessment is undertaken, the information in Chapter 5 on cyanide species reactivity, especially solid–water partitioning, will be useful.

18.3 INTERNATIONAL REGULATORY STANDARDS AND GUIDELINES FOR CYANIDE IN WATER AND SOIL Regulations paralleling those of the United States have been developed in most other industrialized countries for managing discharges of contaminants to water and land, for establishing acceptable concentrations of contaminants in drinking water and surface water, and for guiding cleanup of

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TABLE 18.10 Selected Sediment Quality Guidelines for Cyanide for Specific Applications in the United States Cyanide sediment quality guidelinea (mg/kg) 0.025

0.1 0.175

0.25 0.1

Application Chronic marine sediment threshold (determined with equilibrium partitioning approach) Nonpolluted sediment: USEPA Region 5 harbor classification Moderately-polluted sediment: USEPA Region 5 harbor classification Heavily-polluted sediment: USEPA Region 5 harbor classification Sediment threshold (determined with equilibrium partitioning approach)

Reference [42]

[20,43] [20,43]

[20,43] [44]

a Cyanide species not specified.

contaminated soil and groundwater. Cyanide is and has been a commonly used chemical in all industrialized countries, and thus is a frequently regulated contaminant. In this section, some regulatory standards and guidelines that have been developed in selected industrial countries for cyanide in water and soil are briefly examined.

18.3.1 WATER 18.3.1.1 Drinking Water Maximum acceptable concentrations (MAC) for free cyanide in drinking water that have been established by the World Health Organization (WHO) and various countries around the world are listed in Table 18.11. As seen there, the MAC for free cyanide in Canada, 200 µg/l, is the same as the MCL in the United States. The remaining MAC values listed in Table 18.11 are all less than 200 µg/l, with a low value of 50 µg/l used by Denmark, Germany, and the United Kingdom. The European Union drinking water guideline is 50 µg/l, while that of the WHO is 70 µg/l. 18.3.1.2 Surface Water It appears that relatively few countries have established surface water quality criteria for cyanide for protection of aquatic life. Free cyanide surface water quality criteria developed by some countries for free cyanide in freshwater systems are given in Table 18.12. As seen there, the freshwater cyanide criteria range from 5 to 250 µg/l, with no differentiation of acute or chronic exposure. The development of water quality criteria for aquatic ecosystem protection is in early stages in a number of countries, and especially throughout Europe, where nations are working to develop specific plans to implement the Water Framework Directive of the European Commission that was passed in 2000 [21]. The EC Directive lays out a detailed framework that calls upon participating countries to inventory and characterize surface waters, to assess impacts to these surface waters, and

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TABLE 18.11 Selected International Drinking Water Quality Guidelines for Cyanide

Country or organization Australia Canada Denmark European Union Germany United Kingdom World Health Organization

Max. acceptable conc.a (µg/l)

Reference

80 200 50 50 50 50 70

[45,46] [47] [48] [46] [46] [46] [49]

a As free cyanide (assumed; cyanide species and analytical method not specified in all cases).

TABLE 18.12 Selected International Quality Guidelines for Cyanide in Surface Freshwater (Protection of Aquatic Life) Country

Guideline designation

Free cyanide (µg/l)

Reference

Australia Canada France France France

Freshwater Freshwater Source/soil definition value Fixed impact value: sensitive use Fixed impact value: nonsensitive use

5 30 25 50 250

[45] [47] [50] [50] [50]

to develop plans to address the impacts. Cyanide is among the hazardous substances that will be examined in these evaluations, but specific criteria have not yet been established in most cases.

18.3.1.3 Groundwater Like the United States, other industrial countries have developed regulations and guidelines pertaining to remediation of contaminated soil and groundwater. With respect to groundwater, guidelines have been established to help define when a groundwater has been contaminated and further investigation should be conducted, and to facilitate the establishment of cleanup goals. Table 18.13 lists screening/assessment and action/intervention concentration values that have been established by several countries for different forms of cyanide. The concentration values listed are in the same range as the drinking water MAC values of Table 18.11, with some substantially lower and higher values, interestingly enough. More groundwater standards are forthcoming in Europe, as the European Commission has initiated an effort for a systematic approach to groundwater protection [22].

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TABLE 18.13 Selected International Quality Guidelines for Cyanide in Groundwater

Country

Guideline designation

Australia Austria

Investigation level Groundwater criteria: screening value Groundwater criteria: intervention value Groundwater quality criteria Source/soil definition value Fixed impact value: sensitive use Fixed impact value: nonsensitive use Seepage water quality Target value Intervention value

Austria Denmark France France France Germany Netherlands Netherlands

Total cyanide (µg/l)

Free cyanide (µg/l)

Complex cyanide (µg/l)

naa 30

5 na

na na

[45] [48]

50

na

na

[48]

50

na

na

[48]

na

25

na

[50]

na

50

na

[50]

na

250

na

[50]

50 na na

10b 5.0 1500

na 10c 10d 1500c 1500d

[51] [52,53] [52,53]

Reference

a na: not applicable. b easily-liberated cyanide. c for pH < 5. d for pH ≥ 5.

18.3.2 SOIL Many industrialized countries have established soil quality guidelines, primarily in regard to addressing land contamination. Soil quality criteria for forward-looking protection measures are generally only in the planning and discussion stages. In soil quality guidelines that have been established to facilitate screening/assessment and remediation of contaminated lands, cyanide is frequently included in the lists of contaminants addressed. Table 18.14 provides soil quality guidelines that have been established in selected countries for the purpose of screening assessments at contaminated sites and to guide development of cleanup goals. As is evident in Table 18.14, guidelines are generally aimed at particular land redevelopment scenarios. It is also interesting to note that different forms of cyanide are commonly considered in these guidelines. There is a fairly wide range of acceptable total, free, and complexed-cyanide concentrations represented in the table.

18.3.3 SEDIMENT As noted in the discussion of the U.S. regulations related to contaminated sediments, cyanide is seldom a contaminant of focus in sediment remediation efforts because the most toxic form, free cyanide, is readily biodegradable, and because cyanide species have low bioaccumulation potential. No international sediment quality guidelines or criteria for cyanide were identified in a literature search. As in the United States, most sediment quality criteria developed to date focus on metals and persistent organic compounds [23,24]. Development of sediment quality criteria is expanding

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TABLE 18.14 Selected International Quality Guidelines for Cyanide in Soil

Country

Guideline designation

Australia

Investigation level: residential with garden/accessible soil Investigation level: residential with minimal opportunities for soil access Investigation level: parks, recreational open space Investigation level: commercial/industrial Soil criteria: screening value Soil criteria: intervention value Agricultural land use Residential/parkland use Commercial land use Industrial land use Soil quality criteria Source/soil definition value Fixed impact value: sensitive use Fixed impact value: nonsensitive use Intervention value: playgrounds, housing areas, parks Intervention value: industry and trade areas Target value Intervention value ICRCL trigger value (threshold) for former coal carbonization sites: domestic gardens, allotments ICRCL trigger value (action) for former coal carbonization sites: domestic gardens, allotments ICRCL trigger value (threshold) for former coal carbonization sites: landscaped areas ICRCL trigger value (action) for former coal carbonization sites: landscaped areas ICRCL trigger value (threshold) for former coal carbonization sites: buildings, hard cover ICRCL trigger value (action) for former coal carbonization sites: buildings, hard cover

Australia

Australia Australia Austria Austria Canada Canada Canada Canada Denmark France France France Germany Germany Netherlands Netherlands United Kingdom

United Kingdom

United Kingdom

United Kingdom

United Kingdom

United Kingdom

Total cyanide (mg/kg)

Free cyanide (mg/kg)

Complex cyanide (mg/kg)

naa

250

500

[45]

na

1000

2000

[45]

na

500

1000

[45]

na

1250

2500

[45]

25 500 na na na na 500 na na na 50

na na 0.9 0.9 8.0 8.0 10b 25 50 100 na

na na na na na na na na na na na

[48] [48] [47] [47] [47] [47] [48] [50] [50] [50] [51]

100

na

na

[51]

na na na

1.0 20 25

5.0c 5.0d 650c 50d 250

[52,53] [52,53] [54]

na

500

1000

[54]

na

25

250

[54]

na

500

5000

[54]

na

100

250

[54]

na

500

nld

[54]

a na: not applicable. b acid volatile cyanide. c for pH < 5. d for pH ≥ 5. e nl: no limit specified.

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in Europe, as the European Commission Water Framework Directive [21] calls for expanded assessments of contaminated sediments [25]. At present, in the United States and elsewhere, development of a sediment quality goal with respect to cyanide for a particular situation, if desired, will require a site-specific risk assessment.

18.4 TECHNICAL/REGULATORY ISSUES A fundamental difficulty with the regulatory framework for cyanide is that in many cases regulations are not based on specific cyanide species, but rather in terms of whatever the specified analytical method reports. The use of these operational definitions creates ambiguity when managing actual risks associated with the presence of cyanide species in wastewaters, surface waters, soil, sediment, and groundwater. As a result of this inconsistency between the regulatory definitions and analytical methodologies, there is still much confusion in using different types of cyanide measurements in regulatory compliance contexts.

18.4.1 LACK OF CONSISTENCY: ANALYTICAL METHODS AND REGULATIONS A review of the regulations and guidance for cyanide that were presented in this chapter reveals that the degree of specificity in the regulations and guidance with respect to particular cyanide compounds and chemical forms is often inconsistent with available approved analytical methodologies. That is to say, many more chemical forms are recognized as being of importance than current analytical methods can accommodate. Such a state of affairs lends itself to the often-noted appearance that there is little consistency among various regulatory agencies regarding water quality criteria for cyanide. As MacFarlane et al. [26] have noted, in spite of the stated intention on the part of the USEPA to recognize the difference between toxic and relatively nontoxic forms of cyanide, there is considerable confusion within the regulated community over how current standards for cyanide are to be interpreted, particularly with respect to the analytical methods that are to be used. For example, when the USEPA revised the national primary drinking water regulations in 1992 [27], several comments were received on the proposed MCL and MCLG (maximum contaminant level goal) for cyanide. The concern of the commenters was that while the proposed drinking water maximum contaminant level was specified to be free cyanide, the analytical methods listed imply that total cyanide will be regulated. The USEPA responded [27] by reaffirming that the MCLG and MCL for cyanide apply only to free cyanide. With respect to the analytical methods issue, USEPA stated the following: “USEPA is specifying the use of ‘cyanide amenable to chlorination’ test for determining the ‘free cyanide’ concentrations, while the ‘total cyanide’ analytical technique is being allowed to screen samples. If the ‘total cyanide’ results are greater than the MCL, then the analysis for free cyanide would be required to determine whether there is an exceedance of the MCL.” This, however, does not resolve the confusion, as the cyanide-amenable-to-chlorination measurement includes certain weakacid-dissociable metal–cyanide complexes in addition to free cyanide (HCN, CN− ). The USEPA has had to address comments on the inconsistencies between cyanide regulations and analytical methods in contexts other than drinking water ([28]). It is not uncommon for unofficial, but more appropriate, methods to be approved for use in specific instances (e.g., for the specific discharge of a less toxic form of cyanide), or for variances from permitted discharge levels to be granted in cases where the critical organism upon which the standard is based can be demonstrated to not be present in the receiving water [2,29]. These initiatives point to the importance of cyanide speciation and the insufficient USEPA-approved analytical methods as they relate to determining regulatory compliance.

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18.4.2 CYANIDE TRANSFORMATION IN THE ENVIRONMENT The inconsistency between the regulations and guidelines and available analytical methods is further exacerbated by uncertainties in the fate of cyanide in wastewaters and receiving waters, that is, the role of photolysis with respect to some metal–cyanide complexes, of biodegradation and volatilization of hydrogen cyanide, of synthesis of cyanides by bacteria and other organisms, of the partitioning of cyanide species to solid phases, and of the generation of cyanide by some inorganic and organic compounds upon chlorination. Present water quality criteria for cyanide explicitly or implicitly recognize the importance of these complexities, but are limited in their applicability by the lack of approved evaluation protocols and analytical methods capable of distinguishing among cyanide chemical species

18.4.3 CYANIDE TOXICOLOGICAL DATABASE Many regulations for cyanide in water and soil are in principle focused on protection of human health risk. This has been a substantial challenge for regulatory agencies, however, as cyanide toxicity and availability is strongly dependent on chemical speciation, but conventional analytical techniques for cyanide have provided limited insight into speciation, and information of limited usefulness with respect to cyanide-species availability for some kinds of exposure scenarios. The problems and challenges that exist in the regulation of cyanide species in water and soil based on human health risk continue today (see also Chapter 16). This issue is highlighted here by presenting an example case study that shows the problematic nature of one state-prescribed toxicological database. An example is also provided to describe how certain state-specific solid sample extraction techniques have been developed to address particular cyanide exposure scenarios for soil screening assessments. The first example illustrates the use of inappropriate toxicological information and is related to the State of Michigan toxicological database for evaluating generic clean-up criteria and screening levels for soils. Michigan Department of Environmental Quality (MDEQ) prescribes a oral reference dose (Rfd) of 0.0054 mg/kg-day (measured as available cyanide) [30] for calculation of soil screening levels for residential and commercial purposes. This number is actually based on an acute oral dose of 0.0054 mg/kg that was earlier developed by the Michigan DEQ from a LOAEL (lowest observed adverse effect level) of 0.54 mg/kg, which was obtained from a study of a single human suicide victim who was exposed to cyanide poisoning [31]. This number is very conservative and lacks reliability and accuracy considering the fact that (i) USEPA uses a chronic oral reference dose of 0.02 mg/kgday that was developed using a NOAEL (no observed adverse effect level) of 10.8 mg/kg-day. This NOAEL is based on a two-year dietary study where rats were fed with food fumigated by HCN [32]; (ii) the LOAEL of 0.54 mg/kg is based on a single human victim and hence lacks statistical significance; and (iii) the determination of the absorbed dose of 0.54 mg HCN/kg of body weight in the human suicide victim was not based on analysis of uniform samples of each organ and tissue, but rather on an empirical correlation that is not likely to represent accurately the actual absorbed dose. Further discussion of the Gettler and Baine [31] work is provided in Chapter 16. In summary, given the unreliable nature of the prescribed toxicological database, one could very well overestimate the risk caused by exposure to cyanide-contaminated soils under various environmental settings. The second example relates to the confusion regarding the application of appropriate soil sample preparation (soil sample extraction methodology) for regulatory compliance evaluations. Soil samples submitted for cyanide analysis to a laboratory often are subjected to caustic extraction using USEPA Method 9013 [33] (soil/solid extraction techniques are examined in detail in Chapter 8). However, the extremely high pH conditions involved in caustic extraction solubilizes all forms of cyanide and therefore might not be representative of an ingestion or a dermal mode of exposure. For example, ingestion followed by gut absorption in humans occurs at a pH of 1.5 to 2, while dermal exposure to soil during an excavation scenario may occur predominantly under neutral to moderately acidic conditions (5 < pH < 8). Under such pH conditions, the solubility of iron–cyanide solids

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often present in cyanide-contaminated soils is limited [34–36]. Solubilization of the iron–cyanide solids under these exposure conditions is not well simulated by an alkaline extraction method such as USEPA Method 9013. To address this discrepancy, MDEQ recommends using a neutral leach protocol [37] to simulate direct contact scenarios for cyanide.

18.4.4 CYANIDE AS A CERCLA HAZARDOUS SUBSTANCE As discussed earlier in this chapter, cyanides are regulated as hazardous substances under CERCLA on the basis of application Clean Water Act regulations as well as RCRA regulations. Since the 1970s USEPA has regulated “cyanides” as toxics under Section 307(a) of the Clean Water Act. The scope and meaning of that term has been an issue in at least one case, Commonwealth of Massachusetts v. Blackstone Valley Electric Co. 67 F.3d 981 (1st Cir. 1995), in which the Commonwealth urged USEPA to expand its interpretation of “cyanides” beyond compounds with a freely dissociable cyanide ion. The central issue in the Blackstone Valley case was whether or not ferric–ferrocyanide (FFC), a common, solid-phase cyanide compound encountered at manufactured gas plant (MGP) facilities, is a hazardous substance under CERCLA. In the Blackstone case, Massachusetts was unsuccessful in its cost-recovery action against Blackstone Valley Electric Co. for remediation of an MGP site waste containing FFC. Blackstone defended on the ground that FFC was not one of the cyanides on the CWA toxic pollutant list at the time of the remediation. As a result, USEPA was directed by the First Circuit Court of Appeals to determine whether FFC was included within the term “cyanides” under Section 307(a) of the Clean Water Act. The term “cyanides” appears as a Clean Water Act (CWA) pollutant as Item No. 23 in the list of toxic pollutants pursuant to Section 307(a)(1) of the CWA (40 CFR, 401.15). On January 25, 2001, the USEPA published in the Federal Register its Preliminary Administrative Determination (“PAD”) soliciting comments regarding the USEPA preliminary conclusion that FFC should be considered one of the “cyanides” on the USEPA CWA list. The PAD was published in the Federal Register [38]. After two rounds of notice and comment and after an external peer review, the USEPA issued its Final Administrative Decision on October 6, 2003 [39]. The FAD concluded that FFC fell within the definition of “cyanides.” The USEPA technical basis for categorizing FFC within the meaning of the term “cyanides” was based on a two-step free cyanide release mechanism, namely, FFC solubilization under neutral to alkaline pH conditions into ferrocyanide, which can further dissociate to form toxic free cyanide in presence of sunlight in the environment. However, due to the limited scientific data to support this phenomenon under natural environmental conditions and the fact that the U.S. Food and Drug Administration has prescribed Prussian Blue (or FFC) as an antidote for treatment of thallium or radioactive cesium contamination in humans [40], this argument, has raised criticism. The USEPA final administrative determination on FFC does not change any existing regulations under CERCLA or CWA.

18.5 SUMMARY AND CONCLUSIONS • The majority of U.S. regulations for cyanide in water and soil are based on free cyanide (HCN, CN− ), the most toxic form of cyanide. However, analytical methods used for compliance are often based on “total cyanide” or other aggregate cyanide species measurements. This has led to confusion in the interpretation and use of water and soil regulations for cyanide. • The current use of the term “cyanide(s)” in many water and soil quality regulations in the United States and elsewhere is confusing by not defining the specific cyanide species being regulated.

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• The maximum acceptable concentration of free cyanide in drinking water in the United States and Canada is 200 µg/l. The World Health Organization drinking water guideline for free cyanide is 70 µg/l, while that of the European Union is 50 µg/l. • The freshwater chronic and acute ambient water quality criteria for free cyanide in the United States, for protection of aquatic life, are 5.2 and 22 µg/l, respectively. The saltwater criterion for free cyanide is 1.0 µg/l for both chronic and acute toxicity. The saltwater criterion is being re-examined. Ambient water quality criteria established by other countries for free cyanide are generally in the same range as the U.S. criteria. • The United States and other countries have established industry-specific effluent limitation guidelines, and cyanide is included in these guidelines for many industries, especially chemical manufacturing, metal manufacturing, and metal finishing. • In the United States and elsewhere, groundwater and soil quality guidelines and standards have been developed in the context of contaminated land remediation. Forms of cyanide are commonly detected at contaminated sites. As national groundwater and soil quality standards have not been developed in the United States, individual states have been active in developing such guidelines and standards. Many different approaches and standards have been developed by the states, reflecting the unique conditions and uses of groundwater and soils in the various states. • Few sediment quality criteria for cyanide have been developed in the United States or elsewhere. Cyanide is seldom a contaminant of focus in sediment remediation efforts because the most toxic form, free cyanide, is readily biodegradable, and because cyanide species have low bioaccumulation potential. • The presence of cyanide in solid waste materials is typically assessed by leach testing. USEPA has withdrawn the “reactive cyanide” method for determining reactivity of a cyanide-bearing waste material. • The iron–cyanide solid ferric ferrocyanide (Prussian Blue), which is a common contaminant at former manufactured gas plant sites, has been determined by the USEPA to be included under “cyanides” the Clean Water Act.

REFERENCES 1. ATSDR, Toxicological profile for cyanide (update), U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry, Atlanta, GA, 1997. 2. Brix, K.V., Cardwell, R.D., Henderson, D.G., and Marsden, A.R., Site-specific marine water-quality criterion for cyanide, Environ. Toxicol. Chem., 19, 2323, 2000. 3. USEPA, Methodology for deriving ambient water quality criteria for the protection of human health, EPA-822-B-00-004, U.S. Environmental Protection Agency, Office of Water, Washington, DC, 2000. 4. USEPA, National recommended water quality criteria for the protection of human health, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 68, 75507, 2003. 5. APHA/AWWA/WEF, Method 4500-CN: Cyanide, in Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S. Greenberg, A.E., and Eaton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 6. ASTM, Designation D 2036-98A: Standard test methods for cyanides in water: method A. Total cyanides after distillation, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 7. USEPA, Method 335.2: Cyanide, total (titrimetric, spectrophotometric), Rev. 1980, Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, http://www.nemi.gov, 1979. 8. USGS, Method I-3300-85: Cyanide, total, colorimetric, pyridine–pyrazolone, USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, http://www.nemi.gov, 1985.

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9. USEPA, Method 335.3: Cyanide, total (colorimetric, automated UV), Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, http://www.nemi.gov, 1979. 10. USGS, Method I-4302-85: Cyanide, total recoverable; colorimetric, barbituric acid, automatedsegmented flow, USGS Methods, Vol. A1, U.S. Geological Survey, Denver, CO, http://www.nemi.gov, 1985. 11. USEPA, Method 335.1: Cyanides amenable to chlorination (titrimetric, spectrophotometric), Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, http://www.nemi.gov, 1979. 12. ASTM, Designation D 2036-98B: Standard test methods for cyanides in water: Method B. Cyanides amenable to chlorination by difference, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 13. USEPA, Method OIA-1677: Available cyanide by flow injection, ligand exchange and amperometry, EPA-821/R-99-013, U.S. Environmental Protection Agency, Office of Water, Washington, DC, 1999. 14. USEPA, Common chemicals found at Superfund sites, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, http://www.epa.gov/superfund/resources/chemicals.htm, accessed: March 22, 2005. 15. CLCOM, Cleanup levels for hazardous waste sites, Cleanuplevels.com, http://www.cleanuplevels.com, accessed: April 29, 2005. 16. USEPA, Record of decision, remedial alternative selection: McAdoo Associates site, Schuylkill County, Pennsylvania, EPA-ROD/RO3-85-012, U.S. Environmental Protection Agency, Region 3, Philadelphia, PA, 1985. 17. USEPA, EPA’s contaminated sediment management strategy, EPA-823-R-98-001, U.S. Environmental Protection Agency, Office of Water, 1998. 18. USEPA, The incidence and severity of sediment contamination in surface waters of the United States, EPA-823-R-04-007, U.S. Environmental Protection Agency, Office of Science and Technology, Washington, DC, 2004. 19. Fuhrman, R.L., Memorandum for commanders, major subordinate commands: use of sediment quality guidelines (SQGs) in dredged material management decisionmaking, U.S. Army Corps of Engineers, Washington, DC, October 28, 1998. 20. USEPA, Guidelines for the pollution classification of Great Lakes harbor sediments, U.S. Environmental Protection Agency, Region 5, Chicago, IL (as cited in SAIC, 1991-Ref. [43]), 1977. 21. EC, Directive 2000/60/EC of the European Parliament and of the Council of 23 October 2000 establishing a framework for Community action in the field of water policy, Commission of the European Communities, Brussels, Belgium, 2000. 22. EC, Proposal for a Directive of the European Parliament and Council on the protection of groundwater against pollution, Commission of the European Communities, Brussels, Belgium, 2003. 23. Ahlf, W., Hollert, H., Neumann-Hensel, H., and Ricking, M., A guidance for the assessment and evaluation of sediment quality: a German approach based on ecotoxicological and chemical measurements, J. Soils Sediments, 2, 37, 2002. 24. Burton, G.A., Sediment quality criteria around the world, Limnology, 3, 65, 2002. 25. Brils, J., Sediment monitoring under the EU Water Framework Directive, J. Soils Sediments, 4, 72, 2004. 26. MacFarlane, I.D., Logan, C.M., and Elserod, H.J., Evolving regulatory and risk assessment approaches for cyanide in wastewater discharges, receiving waters, and groundwater, in Proceedings of WEFTEC 97, 70th Annual Conference of Water Environment Federation, Symposium on Remediation of Soil and Groundwater, Chicago, IL, 1997, p. 119. 27. USEPA, 40 CFR Parts 141 and 142, National primary drinking water regulations; Synthetic organic chemicals and inorganic chemicals; Final rule, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 57, 31776, 1992. 28. USEPA, 40 CFR Part 136, Guidelines for establishing test procedures for the analysis of pollutants; Available cyanide in water, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 64, 73414, 1999.

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29. Elserod, H.J., Firstenberg, C.E., Quinn, L.C., Logan, C.M., and Stine, J.F., Water-quality-based permitting for cyanide, in Proceedings of WEFTEC 94, 67th Annual Conference of the Water Environment Federation, Chicago, IL, 1994, p. 309. 30. MDEQ, Table 4. Toxicological and chemical–physical data for Part 201 generic cleanup criteria and screening levels; Part 213 tier I risk-based screening levels, Michigan Department of Environmental Quality, Remediation and Redevelopment Division, http://www.deq.state.mi.us/documents/deq-rrdOpMemo_1-Attachment1Table4ChemicalPhysical.pdf, accessed: April 29, 2005. 31. Gettler, A.O. and Baine, J.O., The toxicology of cyanide, Am. J. Med. Sci., 195, 182, 1938. 32. Howard, J.W. and Hanzel, R.F., Chronic toxicity for rats of food treated with hydrogen cyanide, J. Agric. Food Chem., 3, 325, 1955. 33. USEPA, Method 9013A: Cyanide extraction procedure for solids and oils, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 1, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 2004. 34. Meeussen, J.L., Keizer, M.G., van Riemsdijk, W.H., and de Haan, F.A.M., Solubility of cyanide in contaminated soils, J. Environ. Qual., 23, 785, 1994. 35. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 36. Ghosh, R.S., Nakles, D.V., Murarka, I., and Neuhauser, E.F., Cyanide speciation in soil and groundwater at manufactured gas plant (MGP) sites, Environ. Eng. Sci., 21, 752, 2004. 37. ASTM, Designation D 3987-85(2004). Standard test method for shake extraction of solid waste with water, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2004. 38. USEPA, Preliminary administrative determination document on the question of whether ferric ferrocyanide is one of the “cyanides” within the meaning of the list of toxic pollutants under the Clean Water Act, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 66, 7759, 2001. 39. USEPA, Final administrative determination document on the question of whether ferric ferrocyanide is one of the “cyanides” within the meaning of the list of toxic pollutants under the Clean Water Act, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 68, 57690, 2003. 40. USFDA, Prussian blue for treatment of internal contamination with thallium or radioactive cesium, U.S. Food and Drug Administration, Washington, DC, Fed. Regist., 68, 5645, 2003. 41. USEPA, 40 CFR Part 420, Effluent limitations guidelines, pretreatment standards, and new source performance standards for the iron and steel manufacturing point source category; Final rule, U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 67, 64262, 2002. 42. Bolton, H.S., Breteler, R.J., Vigon, B.W., Scanlon, J.A., and Clark, S.L., National perspective on sediment quality, report by Battelle Memorial Institute for U.S. Environmental Protection Agency (Contract No. 68-01-6986), Washington, DC, 1985. 43. SAIC, Draft compilation of sediment quality guidelines for EPA Region 5 inventory of contaminated sediment sites, Science Applications International Corp., Chicago, IL, 1991. 44. Zarba, C., National perspective on sediment quality, in Contaminated Marine Sediments: Assessment and Remediation, National Academy Press, Washington, DC, 1989, p. 38. 45. Sanaterre, Australian National Environmental Protection Council: Guidelines on the investigation levels for soil and groundwater, Sanaterre Environmental, http://www.sanaterre.com/guidelines/australian.htm, accessed: April 26, 2005. 46. NDWC, International drinking water regulations: the developed world sets the standards, National Drinking Water Clearinghouse, West Virginia University, Morgantown, WV, http://www.nesc.wvu.edu/ndwc/articles/OT/SP03/ Inter_DWRegs.html, accessed: April 25, 2005. 47. EnvCanada, Summary of existing Canadian environmental quality guidelines, Environment Canada, http://www.ec.gc.ca/CEQG-RCQE/English/Ceqg/Water/default.cfm, accessed: April 26, 2005. 48. DEPA, Groundwater protection in selected countries, Danish Environmental Protection Agency, http://www.mst.dk/udgiv/Publications/2002/87-7972-025-0/html/default_eng.htm, accessed: April 26, 2005. 49. WHO, WHO Guidelines for Drinking Water Quality, 3rd ed., World Health Organization, Geneva, Switzerland, 2004.

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50. Sanaterre, French guideline values used for contaminated land management, Sanaterre Environmental, http://www.sanaterre.com/guidelines/french.htm, accessed: April 26, 2005. 51. Mansfeldt, T., Ruhr University — Bochum, Germany, personal communication, 2005. 52. Swartjes, F.A., Risk-based assessment of soil and groundwater quality in the Netherlands: standards and remediation urgency, Risk Anal., 19, 1235, 1999. 53. Sanaterre, The Dutch target and intervention values, Sanaterre Environmental, http://www.sanaterre.com/ guidelines/dutch.htm, accessed: April 26, 2005. 54. Sanaterre, United Kingdom ICRCL guidelines, Sanaterre Environmental, http://www.sanaterre.com/ guidelines/icrcl.htm, accessed: April 26, 2005.

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Treatment Technology: 19 Cyanide Overview George M. Wong-Chong, Rajat S. Ghosh, and David A. Dzombak CONTENTS 19.1

Technology Selection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.1 Waste Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.2 Cyanide Content of the Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.3 Waste Matrix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.4 Other Constituents of Concern . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.5 Treated Waste Quality Requirement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.1.6 Cost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.2 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

387 389 389 389 390 390 390 390 391

Anthropogenic cyanide has long been released to the environment through industrial effluent discharges (e.g., chemical manufacturing, coke plants, gold mining, gas plants, and electroplating effluents) and unregulated disposal of contaminated solid wastes (e.g., aluminum manufacturing spent pot liner, and manufactured gas plants wastes). The latter practice has resulted in contaminated groundwater and soil. Today in the United States, environmental regulations have essentially curbed all uncontrolled discharge practices with industry applying, for the most part, best management practices and best available treatment technologies for the treatment of their wastes. Some areas of earlier waste disposal practice remain problematic (e.g., former gas plant spent oxide box waste and aluminum smelting spent pot liner disposal sites). For the treatment of cyanide in groundwater, wastewater, sludges, and contaminated soil there is available an array of technologies, some of which have been in commercial practice, while others have been extensively evaluated but not used in commercial practice. Table 19.1 presents a list of these technologies, the waste matrix to which they are most applicable, and application status (benchscale, pilot-scale, or commercial practice). The table also provides comments as to industries in which individual technologies are in commercial practice. Fundamental details of these technologies and example commercial applications are presented in subsequent chapters (Chapters 20 to 27).

19.1 TECHNOLOGY SELECTION Selection of an appropriate and economical treatment technology for a given waste depends on a number of factors, including: • Waste chemical and physical characteristics • Quantity of waste to be treated 387

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Carbon adsorption Ion exchange Membrane concentration Air/steam stripping Evaporation Thermal technologies High temperature alkaline chlorination High pressure and temperature alkaline hydrolysis Incineration

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Wet air oxidation

e SPL = spent pot liner.

c Contaminated soil with iron-complexed cyanide. d UV = ultra violet irradiation.

a WAD = weak acid dissociable cyanide. b FeCN = dissolved iron cyanide complexes.































Phytological Chemical oxidation (low temp) Alkaline Chlorination Ozonation w/o UV UVd — Oxidation Other oxidation processes Kastone Air/SO2 w/CO catalyst Separation technologies Precipitation





FeCNb




WADa


Free


Biological Microbial

Technology

Applicable cyanide species

TABLE 19.1 Cyanide Treatment Technology: Overview
































Wastewater
























Groundwater

Waste matrix



















Sludge













Soilc











Bench














Pilot


























Comm

Technology status

Aluminum industry — SPLe treatment Aluminum industry — SPLe treatment Aluminum industry — SPLe treatment

CN recovery in gold mining

Coke plant wastewater with metal-CN; SPLe contaminated groundwater CN recovery in gold mining

Gold mining waste

Electroplating wastewater

Gold mining wastewater; Coke plant wastewater

Comments

388 Cyanide in Water and Soil

Cyanide Treatment Technology

• • • •

389

Waste matrix Other constituents of concern in the waste Treated waste quality required Cost

Brief discussions of these factors in the context of a specific cyanide waste follows. More detailed discussions of these factors in the context of an overall plant- or site-wide waste management strategy are presented in Chapters 26 and 27.

19.1.1 WASTE CHARACTERISTICS In a given cyanide waste, cyanide can exist in numerous forms/species (discussed in Chapters 2, 5, 7, and 8), all of which impact the selection of a treatment technology. These forms include: • Water soluble forms: hydrogen cyanide and CN− ion (measured as free cyanide), weak metal–cyanide complexes (e.g., cyanide complexes with cadmium, copper, nickel, and zinc), and strong metal–cyanide complexes (e.g., cyanide complexes with cobalt and iron). Note that the analytical measurement “weak acid dissociable (WAD) cyanide” includes free cyanide and any weak metal–cyanide complexes, while “total cyanide” includes WAD and any strong metal–cyanide complexes. • Water insoluble forms: transition metal–metal cyanide complex compounds, for example, ferric ferrocyanide or Prussian Blue, Fe4 (Fe(CN)6 )3 (s). When these solid form compounds are placed in water, they dissolve to varying extents, ultimately to their equilibrium solubility limits, yielding low dissolved concentrations of metal–cyanide complexes under natural environmental conditions. This solubility may pose certain treatment challenges and regulatory concerns.

19.1.2 CYANIDE CONTENT OF THE WASTE The cyanide content of the waste (concentration and quantity) and treatment requirements directly impact the selection, size, and cost of the treatment system. For relatively small quantities of waste, as is sometimes the case with contaminated soils, off-site management (e.g., landfill disposal or incineration) would be more cost effective than on-site treatment. Vice versa, with large continuously generated wastewater streams, on-site treatment tends to be more cost effective.

19.1.3 WASTE MATRIX Cyanide occurs in numerous waste matrices, including: • Wastewaters generated from manufacturing operations, which can contain many chemical constituents, and both soluble and insoluble entities. • Contaminated groundwater, which can sometimes contain immiscible organic liquids, also known as “free product” (e.g., light nonaqueous phase liquids, or LNAPL, and dense nonaqueous phase liquids, or DNAPL). • Contaminated soil and soil slurries. • Contaminated sediment (e.g., dredge spoils). • Manufacturing waste sludges. These various waste matrices greatly influence treatment technology selection. In some cases, ancillary treatment processes may be required to ensure the proper operation of the “primary” treatment technology and overall compliance with requisite regulatory limits. The following two

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examples illustrate the differences in treatment schemes needed for similar wastes (e.g., groundwater contaminated with dissolved iron cyanide) in different matrices: 1. Matrix A, groundwater only: effective treatment can be achieved with iron cyanide precipitation only. 2. Matrix B, groundwater with LNAPL: effective treatment requires the separation of the LNAPL prior to iron cyanide precipitation.

19.1.4 OTHER CONSTITUENTS OF CONCERN The presence of other constituents of concern in the waste stream to be treated affects the treatment technology selection and often leads to a train of treatment processes to produce desired treated waste quality. Two examples illustrate the differences in treatment schemes needed for different types of wastes with different regulatory concerns. • Gold mining tailings: These wastes can contain WAD cyanides, strong metal-cyanide complexes, thiocyanate, and trace metals. Biological treatment has proven effective in treating these wastes. The WAD cyanides and thiocyanate are degraded to carbon dioxide, ammonia and sulfate; the strong cyanide complexes and trace metals are adsorbed on to the biological materials; and ammonia is biologically oxidized to nitrate [1]. • Coke plant wastewaters: These wastewaters contain ammonia, WAD cyanide, strong metal-cyanide complexes, phenols, other organics, thiocyanate, sulfide, and trace elements. Treatment of these wastewaters requires a complex treatment train if all of these constituents are of regulatory concern. The treatment train can include: steam stripping to remove ammonia, WAD cyanide and sulfide; biological treatment for ammonia, residual WAD cyanide, phenols, organics, thiocyanate, and sulfide; chemical precipitation for fluoride and strong metal cyanide complexes [2].

19.1.5 TREATED WASTE QUALITY REQUIREMENT This regulatory factor is the primary driving force for treatment and it dictates the degree of treatment that must be achieved. Regulatory requirements and concerns with these requirements are discussed in detail in Chapter 18.

19.1.6 COST The cost of treatment for a given waste stream can vary greatly depending on site-specific conditions and circumstances. Certain publications have provided cost information [1,3–5] specifically focused on treatment of cyanide-bearing wastes. However these cost data are somewhat outdated and should be adjusted for the time since publication. Even with inflation adjustment they should be considered as preliminary guides at best because of the changes in technology markets and global economics. In subsequent chapters, some additional treatment costs and guidelines are provided for specific technologies.

19.2 SUMMARY AND CONCLUSIONS • An array of technologies exists for treatment of cyanide species in wastewaters, groundwaters, soils, and sludges. • Selection of appropriate technology for a given waste depends on various factors, including physico-chemical characteristics of the waste matrix, treatment volume, required treatment level, and cost.

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• Compilations of cost information for certain cyanide treatment technologies exist but are somewhat outdated. Technology cost information from these compilations should be used only for preliminary assessments.

REFERENCES 1. Whitlock, J.L. and Mudder, T., The Homestake wastewater treatment process: biological removal of toxic parameters from cyanidation wastewaters and bioassay effluent evaluation, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd., London, 1998. 2. Wong-Chong, G.M., Sommerfield, F.J., and Velegol, D.J., Coke plant wastewater treatment at Bao Shan Steel, Shanghai, China, in Proceedings of WEFTEC Latin America, Rio de Janiero, 2000. 3. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. 4. Mudder, T., Ed., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. 5. Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Suprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Noyes Datacorp., Park Ridge, NJ, 1988.

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Temperature 20 Ambient Oxidation Technologies for Treatment of Cyanide

Rajat S. Ghosh, Thomas L. Theis, John R. Smith, and George M. Wong-Chong CONTENTS 20.1

Alkaline Chlorination Technologies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.1.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.1.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.1.3 Design Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.1.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.1.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2 Oxidation Technologies with Ozone and Hydrogen Peroxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.3 Design Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3 Photocatalytic Oxidation Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3.1 Process Description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3.3 Design Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4 INCO’s Air/SO2 Process. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.1 Process Description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.3 Design Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.5 Technology Screening Matrix and Additional Technologies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.6 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

394 394 395 396 398 398 398 398 403 403 403 404 404 404 404 405 405 405 406 406 407 408 408 408 408 408 411

Chemical oxidation at ambient temperatures is perhaps the most common treatment technology for cyanide in contaminated waters. Oxidation technologies, such as alkaline chlorination and ozonation perform well for free and weak metal–cyanide complexes (weak acid dissociable cyanide [WAD]) in water, soil slurries, and sludges [1–5]. However, energy-intensive oxidation technologies, such as 393

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ambient temperature photocatalytic oxidation are necessary to treat strong metal– cyanide complexes in water, soil slurries, and sludges [5]. The following ambient temperature oxidation technologies are described in detail in this chapter: • • • •

Ambient temperature alkaline chlorination Ambient temperature oxidation with ozone and hydrogen peroxide Photocatalytic oxidation technologies INCO’s Air/SO2 process

These technologies have been applied for the treatment of water, soil slurries, and sludges containing free cyanide, weak metal–cyanide complexes, or strong metal–cyanide complexes. Descriptions for the technologies follow, and include the following main features: • Process description and implementation • Achievable treatment levels • Design considerations • Critical design conditions • Residuals generated • Technology complexity • Cost information • Status of technology implementation The chapter concludes with a technology summary matrix (Table 20.7) for all the available ambient temperature oxidation technologies.

20.1 ALKALINE CHLORINATION TECHNOLOGIES 20.1.1 PROCESS DESCRIPTION AND IMPLEMENTATION The most widely used technology for the destruction of free cyanide and certain weak metal–cyanide complexes is chlorine oxidation under alkaline conditions, commonly known as alkaline chlorination. Here, free cyanide and certain weakly complexed metal cyanides (i.e., WAD cyanides), such as copper, cadmium, and nickel cyanide, are oxidized to cyanate (CNO− ) and subsequently to carbon dioxide and nitrogen gas. Chlorine gas or hypochlorite (ClO− ) is used as the oxidant, and an alkali (e.g., sodium hydroxide or lime) is used to produce the pH conditions above 9.5 needed to sustain the oxidation reaction. When chlorine gas is used as the oxidizing agent, the process chemistry is given by the following reaction [1,6,7]: CN− + 2NaOH + Cl2 → CNO− + 2Na+ + 2Cl− + H2 O

(20.1)

The above reaction proceeds at significant rates under alkaline conditions (pH 10 and higher) [8]. Addition of alkali is essential to maintain the proper reaction pH and to prevent the generation of any toxic cyanogen chloride (CNCl) or HCN gas, which forms at pH < 10 [6]. The oxidation of cyanide to cyanate is rapid, requiring about 15 to 30 min of contact time and Cl/CN dose of about 3 (on a mass basis). The complete destruction of cyanide can be accomplished by lowering the pH of the solution after cyanate formation to 9 and addition of excess chlorine. This second reaction proceeds as follows [7]: 3Cl2 + 2CNO− + 4NaOH → 2CO2 + N2 + 2Cl− + 4Na+ + 4Cl− + 2H2 O

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TABLE 20.1 Typical Operating Conditions for a Two-Stage Alkaline Chlorination Process Stage 1 2

pH

Chlorine dose (g Cl/g CN)

NaOH dose (g NaOH/g CN)

Redox potential (mV)

Retention time (min)

9.5–11 8.0–8.5

2.7–3.0 4.1–4.5

3.1–3.4 4.2–4.6

350–400 600

30–60 30–60

Source: Data from Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Surprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Corp, N.D., Ed., Noyes Data Corp., Park Ridge, NJ, 1998.

In cases where a metal–cyanide species is oxidized, the liberated metal generally forms a hydroxide precipitate under the alkaline conditions of the reaction. Treatment of thiocyanate (SCN− ) by alkaline chlorination occurs in the pH range of 10 to 11.5 according to the following reaction: − 2SCN− + 8Cl2 + 20OH− → 2CNO− + 2SO−2 4 + 16Cl + 10H2 O

(20.3)

The alkaline chlorination process for free and WAD cyanide can be operated as a one- or two-step process in either batch or continuous flow. In the two-step process, the first step is used for oxidation of cyanide to cyanate; in the second step, cyanate is oxidized to carbon dioxide and nitrogen. Cyanate, however, can also be hydrolyzed to CO2 and NH3 by adjusting pH to the 7 to 8 range, which reduces the chlorine demand. There is extensive full-scale application of this technology in electroplating and gold mining operations. Table 20.1 gives typical operating conditions for a two-stage, full-scale continuous flow alkaline chlorination unit for treating free and WAD cyanide. Figure 20.1 presents a schematic flow diagram of a typical alkaline chlorination system for the treatment of cyanide in tailings pond decant water [9]. Although the figure shows chlorine gas being used, this can be replaced by hypochlorite solution, which would eliminate the recirculation pump and chlorine eductor; however, a hypochlorite solution feed pump would still be required. The hypochlorite feed pump or chlorine gas feed would be oxidation–reduction potential (ORP) controlled and the lime/alkaline feed would be pH controlled. Tables 20.2 and 20.3 present operating parameters and effluent quality produced by alkaline chlorination systems at four gold mining operations. It should be noted that residual chlorine is toxic to many species in the environment and discharge of effluents with high residual chlorine concentrations can be problematic and, in some instances, will be prohibited. For the treatment of certain weak metal–cyanide and strong metal–cyanide complexes, modifications to this process are implemented, including increasing the temperature and retention times in the reaction vessel [6,10,11]. Details of high temperature alkaline chlorination technology are provided under thermal technologies in Chapter 22.

20.1.2 ACHIEVABLE TREATMENT LEVELS Weakly complexed metal cyanides are typically reduced to a concentration less than 1 mg/l, while free cyanide concentrations following alkaline chlorination are usually less than 0.2 mg/l. These performance levels will depend on chlorine dosage, reaction pH, reaction time, and the general chlorine demand of the waste. This technology is not applicable for strongly complexed metal cyanides like iron– or cobalt–cyanide complexes.

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Chlorine gas or hypochlorite Lime slurry Barren solution or tailing pond water

pH

ORP

Mixing

Eductor

Reactor tank(s) 0.5–1.5 h pH 10–11.5 Recirculating pump Solid tails

Tailings sump

To tailings pond

FIGURE 20.1 Schematic flow diagram of a typical alkaline chlorination system. (Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.)

TABLE 20.2 Operating Parameters for Full-Scale Alkaline Chlorination Operations Baker mine

Carolin mine

Giant Yellowknife mine

Parameter

Mosquito Creek mine

Mill capacity (Tpd)a Solids cyanided Solid feed rate (Tpd)a Treatment mode Solution treated

100 Ore 100 Batch Barren

100 Ore 100 Cont. Barren

1250 Concentrate 75 Cont. Barren

Solution rate

14.4 m3 /day

216 m3 /day

Form of chlorine

3 to 5.5 m3 batches/day Gaseous

1200 Roaster calcine 140 Cont. Tailings pond overflow 6545 m3 /day

Gaseous

Gaseous

No. reactor tanks Retention time (h) pH pH control Chlorine control

1 6 11 Manual Manual

Calcium hypochlorite 2 14 11.5 Manual Manual

1 8 11 Auto Manual

1 0.5 11.5 Auto Manual

a Tpd = metric tons (tonnes) per day.

Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.

20.1.3 DESIGN CONSIDERATIONS The critical design parameters for alkaline chlorination include chlorine/cyanide (Cl/CN) ratio, reaction pH, and reaction time. The technology is well suited for treatment up to 5000 mg/l of

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TABLE 20.3 Performance Data for Full-Scale Alkaline Chlorination of Gold Mill Effluents Constituents, mg/l Mine

CNaT

CNbW

CNS

Cu

Fe

Ni

Zn

As

NH3

TRCd

Baker Influent Effluent % removal

2000 8.3 99.6

1900 0.7 99.9

1100c — —

290 5.0 98.3

2.4 2.8 —

— — —

740 3.9 99.5

— — —

— — —

— 2800e —

Carolin Influent Effluent % removal

1000 170 83

710 0.95 99.9

1900c — —

97 0.38 99.6

150 53 64.7

— — —

110 5.8 94.7

— — —

— — —

— 190 —

Mosquito Creek Influent Effluent % removal

310 25 91.9

226 0.49 98.8

330c — —

10.0 0.33 96.7

9.4 8.0 14.9

— — —

93 1.4 98.5

— — —

— — —

— 320 —

Giant Yellowknife Influent Effluent % removal Polishing pond O/F % removal

7.5 1.3 82.7 0.15 98

7.1 1.2 85.1 0.09 98.7

6.3 1.0 84.1 — —

6.7 0.09 98.7 0.03 99.6

<0.1 <0.1 — <0.1 —

1.2 0.7 41.7 — —

0.1 0.1 — <0.1 —

12.1 — — 0.14 99.7

— — — 9.4 —

— — — 1.1 —

All samples unfiltered. a CN = total cyanide by distillation. T b CN = weak acid dissociable cyanide by ASTM Method C. W c Analysis not available due to analytical difficulties. d TRC = total residual chlorine. e Additional chlorine added with a view to destroying cyanide contained in solid tailings slurry.

Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.

free cyanide using batch systems, while continuous processes with flow rates up to 5 gpm can treat up to 1000 mg/l, with optimal treatment efficiency usually achievable for concentrations below 100 mg/l and influent flow rates up to 100 gpm [6,7,12]. Waste chlorine demand greatly influences Cl/CN ratio; chlorine demand does not depend only on cyanide content. The technology is not suitable for waste streams containing strong metal–cyanide complexes, such as ferro- or ferricyanide and high concentrations of thiocyanates (SCN− ). Moreover, optimal efficiency is achieved for influents containing less than 100 mg/l of total suspended solids (TSS), less than 1000 mg/l of total dissolved solids (TDS), pH levels between 9 and 13, and ORP greater than 200 mV. As far as residuals are concerned, metal hydroxide sludges can be generated if the influent stream contains appreciable amounts of weak metal–cyanide complexes, or metals in other forms. Weaker complexes that dissociate during the process of oxidation will liberate metal cations, leading to the formation of metal hydroxides under alkaline pH conditions. Residual chlorine and chloramines are also generated, which, because of their toxic nature, should be removed by dechlorination prior to discharge. At pH < 9, generation of CNCl, a toxic gas, as an intermediate during the oxidation of

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cyanide to cyanate is a concern. Careful control of pH and ORP should be in place to prevent any evolution of CNCl gas. The technology is relatively easy to implement and operate. It requires basic wastewater treatment unit operations and continuous monitoring of pH to prevent production of CNCl and HCN. Chlorine gas handling and leakage pose possible health hazards. If metal hydroxide sludges are generated, they may require additional treatment for stabilization prior to disposal. Moreover, the heat of reaction from chlorine and cyanide decomposition may require some form of temperature control before the final effluent can be discharged to the sewer.

20.1.4 COST OF THE TECHNOLOGY Capital costs for a typical 500 gpm system for treating waste streams that contain free and WAD complexes has been reported as approximately $300,000 (1990 cost basis), with typical operation and maintenance (O&M) costs varying between $5 and $7 per kilogram of cyanide destroyed [6,9,12,13].

20.1.5 TECHNOLOGY STATUS Alkaline chlorination is a well-established, commercially practiced technology with many successful full-scale applications in place in electroplating and gold mining industries [6,9,12,13]. Prefabricated chemical feed and monitoring equipment suitable for implementing this technology are commercially available. However, some bench-scale testing for a particular application usually is desirable for determination of optimal Cl/CN dose, pH conditions, and reaction time.

20.2 OXIDATION TECHNOLOGIES WITH OZONE AND HYDROGEN PEROXIDE 20.2.1 PROCESS DESCRIPTION AND IMPLEMENTATION These processes involve the oxidative destruction of free and WAD forms of cyanide by either ozone or hydrogen peroxide under alkaline pH (9–11) conditions. Oxidation of cyanide (CN− ) to cyanate (CNO− ) occurs in 10–15 min in the presence of excess ozone under alkaline conditions (9 < pH < 10) according to the following reaction [14]: CN− + O3 → CNO− + O2

(20.4)

Gurol and Bremen [3] reported a first-order reaction rate coefficient (2600 ± 700 M −1 sec−1 ) for ozonation of free cyanide at pH 11.2. Figure 20.2 presents the observed pseudo-first-order rate constant for ozone decay as a function of total cyanide concentration. As shown in this figure, the cyanide oxidation rate increases with increase in pH. Rate expressions for ozone oxidation of cyanide at three different pH values are as follows [3]: −d[O3 ]/dt = (2600 ± 700)[CNT ]0.55±0.06 [O3 ]

at pH = 11.2

(20.5)

−d[O3 ]/dt = (2700 ± 850)[CNT ]0.83±0.14 [O3 ]

at pH = 9.5

(20.6)

−d[O3 ]/dt = (550 ± 200)[CNT ]1.06±0.1 [O3 ]

at pH = 7.0

(20.7)

The presence of copper was found to catalyze the cyanide oxidation process according to the following reaction [15]: − 2Cu+ + 11CN− + 3O3 → 2Cu(CN)3− 4 + 3CNO + 3O2

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Phosphate solutions 3.0

1 2 3

pH 11.2 pH 9.5 pH 7.0

Log kobs (sec–1)

2.5

2.0 2 1

1.5

1.0

3

0.5

0 –4.0

–3.0

–2.0 Log [CNT], M

–1.0

0

FIGURE 20.2 Observed pseudo-first-order rate constant for ozone decay vs. total cyanide concentration on log scales. (Source: Reprinted with permission from Gurol, M.D. and Bremen, W.H., Environ. Sci. Technol., 19, 804, 1985. Copyright 1985. American Chemical Society.)

Additional discussion of this reaction and the catalytic effect of the copper is provided in Chapter 5. In the presence of excess ozone, cyanate is hydrolyzed to bicarbonate and nitrogen according to the following reaction [14]: 2CNO− + 3O3 + H2 O → 2HCO− 3 + N2 + 3O2

(20.9)

This second stage reaction is much slower than the cyanate formation reaction and is usually carried out in the pH range of 10 to 12 where the reaction rate is relatively constant. Temperature variation within the ambient range does not have a significant effect on the reaction rates. However, the use of ultraviolet (UV) light to enhance radical formation [6] and the presence of copper catalyst [12] have each been shown to increase the rate of the second stage reaction. The metal–cyanide complexes of cadmium, copper, nickel, silver, and zinc are readily oxidized by ozone. For treatment of strong metal–cyanide complexes, such as iron– and cobalt–cyanide, modifications to the existing process are implemented, including prolonged UV light exposure to promote photodissociation [4,5]. However, Gurol and Holden [15] reported oxidation of iron–cyanide complexes in the presence of excess ozone (ozone to iron cyanide ratio of 30:1 on a molar basis) under laboratory conditions. Thiocyanate/SCN− is readily oxidized by ozone [16]. Layne et al. [16] determined that for − pH > 11, SCN− reacts with ozone to form CN− and SO2− 4 , and the free CN is subsequently − oxidized to CNO as shown in reaction (20.4).

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Hydrogen peroxide provides another alternative in treating free and weakly complexed metal cyanides in waters and wastewaters. Although H2 O2 is a weaker oxidizing agent than ozone (standard electrode potential of 0.878 V in alkaline solution compared to 1.24 V for ozone under same solution conditions), cyanide can be fully converted by hydrogen peroxide to ammonia and carbonate under alkaline conditions, according to the following reactions: CN− + H2 O2 → CNO− + H2 O

(20.10)

CNO− + H2 O + OH− → NH3 + CO2− 3

(20.11)

The first reaction is optimal in the pH range of 9.5 to 10.5 [8]. The second reaction, however, is very slow under alkaline condition and increases as pH decreases [17]. The cyanide oxidation rate also depends on the excess hydrogen peroxide concentration, cyanide concentration, and temperature. The reaction rates can also be enhanced by the presence of a metal catalyst, such as copper, which ultimately reacts with ammonia to form a tetraamino copper complex that is largely nonreactive [8]. Copper-catalyzed hydrogen peroxide oxidation of WAD cyanide complexes in wastewater is practiced commonly in the gold mining industry [9]. The destruction of weak metal–cyanide complexes occurs according to the following reactions: − M(CN)−2 4 + 4H2 O2 + 2OH

Cu catalyst −→ 4CNO− + 4H2 O + M(OH)2 (s)

2− CNO− + 2H2 O −→ NH+ 4 + CO3

(20.12) (20.13)

where M is a metal cation, such as Cu or Zn. The copper, which is added as a catalyst or present in 4− the waste as Cu(CN)− 2 , can react with strongly complexed Fe(CN)6 to form an insoluble bimetallic complex according to the following reaction: +2 Fe(CN)4− −→ Cu2 [Fe(CN)6 ](s) 6 + 2Cu

(20.14)

It is customary to add copper sulfate pentahydrate as the catalyst to produce a copper concentration of about 10 to 20% of the WAD cyanide concentration. The peroxide dose needed for successful oxidation of cyanide species may be 200 to 450% of the required amount indicated by stoichiometry [9]. The high peroxide dosage rate is reflective of the presence of other oxidizable materials in the wastewater that can compete for the peroxide, as well as the inherent loss of oxidation capacity as some of the peroxide may decompose to oxygen and water: 2H2 O2 −→ O2 + 2H2 O

(20.15)

To reduce these decomposition losses, peroxide stabilizers such as silicate (employed in Degussa’s SILOX process) and sulfuric acid, which forms peroxymonosulphuric acid (Caro’s acid), have been developed and deployed with substantial savings over the conventional peroxide process [18]. Figure 20.3 presents a schematic flow diagram of a typical hydrogen peroxide treatment system for cyanide [18]. As shown in this figure, hydrogen peroxide is added to the first reaction tank along with the influent solution. In the second mixing tank, copper is added as copper sulfate to catalytically promote the cyanide oxidation reaction. The supernatant from the second mixing tank then goes to the third tank, where enough settling of solid sludges (copper–iron–cyanide solids; iron hydroxides) and increased residence time causes complete removal of cyanide, and cyanide-free supernatant is discharged into the tailings pond. Figure 20.4 and Table 20.4 present a schematic flow diagram and performance data for a continuous tailings slurry treatment system using hydrogen peroxide at the OK Tedi Mine in Papua,

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CuSO4 catalysts (if required) Tailings pulp or Barren solution

To tailings pond

Feed pump H2O2 storage

Reaction tanks

FIGURE 20.3 Schematic flow diagram of a typical hydrogen peroxide treatment system for cyanide. (Source: Botz, M. et al., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

Control system Activator CN Caroate NaOH H2SO4

Control unit

Multiplier

H2O2 pumps

Control stream

pH Redox

Measuring cell

Control valve H2O2

Main tailings stream Tailings slurry 1100 m3/h 110–300 mg /l CN

Flow meter

Reaction tank

Sample for analysis

1–10 mg/l CNT <0.3 mg/l WAD CN

T

FIGURE 20.4 Schematic flow diagram for the Degussa hydrogen peroxide process at the OK Tedi Mine. (Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.)

New Guinea. Because of the lack of suitable means to determine the necessary dosage of H2 O2 quickly and accurately enough to allow efficient use of the reagent for treatment of large effluent flows, a continuous automatic titration is implemented in a small sidestream as depicted in Figure 20.4. The pH of the sidestream is adjusted automatically to a particular value, and a fast-acting strong oxidizing agent is dosed. The rate of dosage is controlled by a redox measurement carried out in the presence of a special catalyst (“Activator CN”). Simultaneous to the addition of the strong oxidizing agent (an aqueous solution of “caroate,” potassium monopersulfate) to the sidestream, H2 O2 , at a concentration of 70% by weight, is added to the main tailings stream via a control valve. The opening

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TABLE 20.4 Tailings Slurry Characteristics after Degussa Hydrogen Peroxide Treatment at OK Tedi Mine Parameter

Before H2 O2 Treatment

After H2 O2 Treatment

Tailings flow Solids content pH Free cyanide WAD cyanide Total cyanide Dissolved Cu Dissolved Zn Dissolved Fe

1100 m3 /h 45% 10.5–11.0 50–100 mg/l 90–200 mg/l 110–300 mg/l 50–100 mg/l 10–30 mg/l 1–3 mg/l

1100 m3 /h 45% 10.2–10.8 Undetectable <0.5 mg/l 1–10 mg/l <0.5 mg/l <0.1 mg/l 1–3 mg/l

Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.

TABLE 20.5 Treatment Performance for Three Hydrogen Peroxide Treatment Plants Before Treatment (mg/l)
CN Case study #1 Pond overflowa Case study #2 Barren bleedb Case study #3 Heap leach solutionc

After Treatment (mg/l)

WAD CN

Cu

Fe


CN

WAD CN

Cu

Fe

19

19

20

<0.1

0.7

0.7

0.4

<0.1

1350

850

478

178

<5

<1

<5

<2

353

322

102

11

0.36

0.36

0.4d

<0.1

a Preliminary plant results from pre-operational test runs. b Typical results during first six months of operation. c Average of 25 measurements made over 10 days of plant operation. d Value dropped from 1.0 to 0.4 over 4 days due to coagulation and settling.

Source: Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.

of this valve is controlled by a signal obtained by multiplying the signal from the control unit by a second signal obtained from a tailings flow meter. Table 20.5 presents performance data from three other hydrogen peroxide treatment facilities at gold mining sites. While the data in Tables 20.4 and 20.5 show excellent removal of cyanide by oxidation and precipitation of metals, it must be recognized that these facilities are only used for treatment of primary constituents of concern, like cyanide. Hydrogen peroxide treatment does not affect ammonia, nitrate, or thiocyanate; treatment of these constituents will require additional treatment units. Hydrogen peroxide oxidation for free cyanide can also be effective under alkaline conditions, and in the presence of a metal catalyst (Fe, Al, Ni) or formaldehyde. The patented Kastone

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Process uses H2 O2 and formaldehyde to oxidize free cyanide to cyanate at 49–54◦ C and at a pH of 10–12 [12].

20.2.2 ACHIEVABLE TREATMENT LEVELS Free and weakly complexed cyanides are typically reduced to a concentration less than 0.1 mg/l depending on ozone or hydrogen peroxide dose, reaction pH, and reaction time. The oxidation of cyanide by ozone and hydrogen peroxide usually occurs rapidly up to cyanate formation. Oxidation of cyanate by ozone, however, is a slow reaction and cyanate may accumulate in the solution until cyanide is completely oxidized. Hydrogen peroxide is a weaker oxidant than ozone and requires greater doses for the same level and rate of cyanide destruction. In addition, with hydrogen peroxide, the cyanate oxidation reaction rate increases with decrease in pH and the presence of copper catalyst. Achievable treatment levels for cyanide using Kastone Process could be as low as 0.1 mg/l.

20.2.3 DESIGN CONSIDERATIONS The critical design parameters include ozone/cyanide (O3 /CN) or H2 O2 /CN ratio, reaction pH, and reaction time. The presence of significant amounts of organic material or reduced inorganic species can significantly increase the ozone or hydrogen peroxide demand. Full scale oxidation systems are usually limited to total cyanide concentrations of less than 40 mg/l and with less than 1% organic matter [6], and are unsuitable for waste streams containing strong metal–cyanide complexes and high thiocyanate content. Optimal waste stream handling conditions are as follows: TSS < 100 mg/l, TDS < 1000 mg/l, and pH of the stream between 5 and 7. Ozonation is usually most economical for flows less than 500 gpm. Moreover, this technology requires a continuous supply of cooling water (typically 4000 l of water per kg of ozone). Similar restrictions are applied for treatment systems using hydrogen peroxide as the oxidant. As far as residuals are concerned, metal hydroxide sludges can be generated if an influent stream contains appreciable amounts of weak metal–cyanide complexes. Moreover, the presence of cyanate in the product stream may require additional treatment prior to discharge. The oxidation technologies involving ozone and hydrogen peroxide are more complex than the alkaline chlorination process. For ozonation, on-site ozone generators, including air compressors and oxygen concentrators, are used in addition to the process reactor, along with their dedicated control systems. Like alkaline chlorination, the technology requires extensive health and safety training for operators, especially when dealing with a strong oxidizer such as ozone. The benefits of using ozone over chlorine are: (i) stronger oxidation potential, (ii) on-site generation resulting in reduced transportation, storage, and handling costs, and (iii) elimination of potential formation of chlorinated organics. However, on-site generation facilities and power requirements may incur significant capital and operating costs [19].

20.2.4 COST OF THE TECHNOLOGY The capital cost of ozone oxidation technology is significantly higher than the alkaline chlorination process. It requires higher initial cost, related primarily to the on-site ozone generation equipment, and the need for a continuous supply of cooling water. Capital costs for a typical 500 gpm ozonation system have been reported as $875,000 (1988 cost basis); typical O&M costs are around $2/kg of cyanide destroyed [6]. The capital and operating costs associated with hydrogen peroxide systems are usually lower than ozonation systems of the same scale, but are higher than conventional alkaline chlorination processes.

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20.2.5 TECHNOLOGY STATUS Ozonation and hydrogen peroxide application are well-established technologies with limited fullscale applications in place [6], mainly in the mining and electroplating industries. Prefabricated chemical feed and monitoring equipment suitable for implementing this technology are commercially available.

20.3 PHOTOCATALYTIC OXIDATION TECHNOLOGY 20.3.1 PROCESS DESCRIPTION This three-step process involves UV-light-aided photodissociation of metal–cyanide complexes, including the strong iron– and cobalt–cyanide complexes, to free cyanide. The liberated free cyanide is further oxidized to CO2 and NO− 3 , using either ozone or H2 O2 in the presence of a TiO2 catalyst. Photodissociation of ferri- and ferrocyanide complexes, discussed in Chapter 5, has been extensively studied in the laboratory, over a wide range of pH conditions, for the purpose of treating waters contaminated with iron–cyanide complexes [5,20–24]. The photocatalytic oxidation reaction scheme for iron–cyanide complexes has been described by Schaefer [22] as follows: − + − Fe(CN)3− 6 + 3H2 O + hν → CN + Fe(OH)3 (s) + 3H + 3e

(20.16)

CN− + oxidant → CNO−

(20.17)

CNO− + oxidant → CO2 + NO− 3

(20.18)

As noted previously, ozone provides much more rapid reaction rates than hydrogen peroxide [6], and the cyanate oxidation reaction is usually slower than the cyanide oxidation and the initial photodissociation reactions. However, UV irradiation in combination with hydrogen peroxide or ozone results in the formation of OH• radicals, which are strong oxidizing agents capable of oxidizing iron–cyanide complexes. Photocatalytic oxidation may be implemented in one or two stages, and in batch or continuous flow mode under conditions of ambient temperature and pressure. In a one-stage system, photodissociation and oxidation occur in the same reactor vessel. In a two-stage system, the first stage is used to photodecompose the iron–cyanide complex under alkaline conditions at a UV wavelength of 350 nm, and the second stage is used for complete oxidation of the free cyanide ion in the presence of an oxidant and a catalyst. Figure 20.5 shows the typical features of a two-stage photocatalytic oxidation system. Note that an intermediate filtration step is performed to remove any metal oxide and hydroxides produced under the alkaline pH conditions from free iron and other metals produced upon photodissociation.

20.3.2 ACHIEVABLE TREATMENT LEVELS Under bench-scale laboratory conditions, Schaefer [22] achieved complete photocatalytic oxidation of an aluminum reduction wastewater stream containing 64 to 85 mg/l of soluble ferrocyanide in 2 h to less than 0.5 mg/l in the effluent. However, complete destruction of cyanide to carbon dioxide did not occur, and the reaction sequence slowed in the second stage (Equation [20.17]) with the formation of cyanate. The first-order rate constant for the dissociation of ferrocyanide at an ozone dose of 865 mg/min was 0.0332 min−1 . To determine the effect of variable ozone dosage, additional experiments performed at a smaller ozone dose of 140 mg/min generated an even lower photodissociation rate of 0.0089 min−1 . Longer reaction time and presence of suspended TiO2 catalysts were identified as possible approaches to improve performance.

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Acid

O3

Effluent

Pre-treatment • Cyanide waste & caustic mixing • Decolorization • Solids separation

Pump

= UV lamp

TiO2

TiO2

Filter Stage 1: photolysis

Stage 2: oxidation

FIGURE 20.5 Two-stage photocatalytic reactor. (Source: Copyright © 1997. Electric Power Research Institute. TR-108596. Technology Review: Treatment of Complexed Cyanide in Water. Reprinted with permission.)

20.3.3 DESIGN CONSIDERATIONS Photocatalytic treatment is usually most economically feasible for small flow rates, that is, less than 25 to 30 gpm, and is most suitable for treating waste streams with the following characteristics: TSS < 100 mg/l, TDS < 200 mg/l, pH > 9, and low soluble iron content. Influent turbidity and production of iron oxide/hydroxides during the treatment process may inhibit UV light penetration and reduce treatment efficiency. This can be overcome using continuous filtration [5,22] or chelating agents such as EDTA to hold the released iron in solution [25]. In addition, the presence of significant amount of organics and inorganics in the waste stream can add significantly to the oxidant demand. Hence, application of UV oxidation technology will usually be limited to relatively clean waters. Prefabricated photocatalytic reactors are available from commercial vendors selling wastewater disinfection technology. However, there is no significant commercial experience with implementation of this technology for treatment of cyanide in water. The technology, if implemented, also needs continuous monitoring and maintenance to prevent sludge buildup and the resultant reduction in photointensity during operation.

20.3.4 COST OF THE TECHNOLOGY The capital costs for a full-scale 25 gpm continuous treatment system that treats influent with cyanide concentration as high as 100 mg/l could range anywhere from $1.4M (UV with H2 O2 ) to $1.83M (UV with O3 ). The inherent operating costs for this technology is on the high end, with operation and maintenance costs ranging between $0.28M/yr (UV with H2 O2 ) to $0.26M/yr (UV with O3 ) (2001 cost basis; Alcoa Inc., internal communication).

20.3.5 TECHNOLOGY STATUS Even though extensively studied in the laboratory, field scale implementation of this technology has been limited. A major advantage of UV/peroxide and UV/ozone oxidation is that no undesirable by-products (e.g., ammonia) are generated. Prefabricated photocatalytic reactors are available from

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commercial vendors. Peroxidation systems, now part of Calgon Carbon Corp., manufactures a modular system comprising a UV light source (200 to 280 nm) and hydrogen peroxide storage and feed equipment. This system has been installed at many locations, though no reports of its use for cyanide treatment have been published.

20.4 INCO’S AIR/SO2 PROCESS 20.4.1 PROCESS DESCRIPTION A patented cyanide oxidation process is the Air/SO2 process [26,27] that was developed by the International Nickel Company of Canada (INCO). The process is similar to other oxidation processes, requiring reaction vessels with mixing to contact the oxidants with cyanide in the wastewater (Figure 20.6). This process utilizes air and SO2 to oxidize free cyanide and weakly-complexed metal cyanides in the presence of a copper catalyst. The process reactions are similar to those for chlorine and hydrogen peroxide in that cyanate is the oxidation product, as shown below: 4CN− + 4SO2 + 4O2 + 4H2 O −→ 4CNO− + 4H2 SO4

Copper sulfate (if required)

(20.19)

Lime

pH

7 10

Tailings slurry or decantate To tailings pond

Sulfur dioxide

SO2 storage vessel

Reactor Retention: 0.3 to 2 h

Air Air blower

FIGURE 20.6 Schematic diagram of the INCO SO2 /Air oxidation process for the removal of cyanide. (Source: Botz, M. et al., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

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Lime is added to the reaction vessel to neutralize the sulfuric acid that is generated. A pH in the range of 7 to 10 is typical. The stoichiometric SO2 requirement is 2.46 g/g of CN− oxidized, but in practice, the actual usage ranges from about 3.5 to 4.5 g SO2 per g of CN− oxidized. The SO2 required in the reaction may be supplied as liquid SO2 or as sodium metabisulfite (Na2 S2 O5 ). Under normal operating conditions, thiocyanate is only partially (10 to 20%) oxidized [9] according to the following reaction: SCN− + 4SO2 + 4O2 + 5H2 O −→ CNO− + 5H2 SO4

(20.20)

During the course of the oxidation, any ferricyanide complex is reduced to ferrocyanide complex, which in turn can react with copper, nickel, or zinc to form a low-solubility precipitate. Excess copper, nickel, or zinc form their respective hydroxide precipitates at a pH of 8 to 10.

20.4.2 ACHIEVABLE TREATMENT LEVELS The INCO Air/SO2 process is generally able to render effluents with total cyanide levels below 1 mg/l, even with influent total cyanide levels as high as 2000 mg/l. Tests performed by INCO using a continuous one-stage reactor showed that with a hydraulic residence time of 97 min, a feed stream containing 1680 mg/l CNT was reduced to 0.13 mg/l total cyanide [28]. Table 20.6 presents performance data for full-scale SO2 /Air oxidation treatment of gold mine tailings slurries, barren solutions, and tailing pond decant waters. These data show relatively good cyanide removal; substantial metals precipitation can also be inferred from the data. However, like the other oxidation processes, SO2 /Air oxidation results in limited thiocyanate treatment (10 to 20%) and no treatment of ammonia and nitrate.

TABLE 20.6 INCO’s Air/SO2 Destruction of Cyanide in CIP Tailings, CIL Tailings, Repulped Tailings, Barren Solution, and Pond Water Cyanide concentration, mg/l Mine Colosseum Ketza River Equity Casa Berardi Weatmin Premiere Golden Bear McBean (barren) Lynngold (pond) Mineral Hill (barren) Lac Shortt (pond) Citadel (barren) St. Andrew (pond)

Reagent usage, g/g CNT

Before feed

After effluent

SO2

Lime

Cu

364 150 175 150 150 205 370 106 350 10 350 15

0.4 5.0a 2.3 1.0 <0.2 0.3 0.2 0.6 0.5 0.5 5.0a 1

4.6 6.0 3.4 4.5 5.8 2.8 4.0 7.0 6.0 5.0 4.0 5.0

0.12 0 0 — — — 4.0 9.0 9.0 — — —

0.04 0.30 0.30 0.10 0.12 — 0 0.12 0 0 0 0.10

a Complete cyanide destruction not required to meet permit levels.

Source: Data from Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991.

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20.4.3 DESIGN CONSIDERATIONS The optimum operating conditions for free cyanide and weak metal–cyanide complexes are pH of approximately 9, cyanide to cupric ion mass ratio of 5:1, and cyanide to sulfur dioxide mass ratio between 1:3 and 1:7 [9]. Commercial units have been successful in treating tailings pulp up to 40% solids at a flow rate of 270 kg CN− /h [29].

20.4.4 COST OF THE TECHNOLOGY Available cost information for the Air/SO2 process is very limited. Using a Canadian Dollar exchange rate of $1.185 per $1 US (for 1989), limited vendor-specific information indicates capital cost in the vicinity of $210,000 (1989 cost basis) for a 1 kilo ton/day tailing slurry treatment system with operating cost in the range of $1.36/ton of tailings treated [9].

20.4.5 TECHNOLOGY STATUS The INCO cyanide destruction technology is proprietary. As of 1998, 45 licenses had been issued worldwide for full-scale applications [30] and over 70 treatment facilities had been installed [9].

20.5 TECHNOLOGY SCREENING MATRIX AND ADDITIONAL TECHNOLOGIES The various ambient temperature oxidation technologies described in this chapter are summarized in Table 20.7. The table includes information on performance, cost, and implementation experience and can be used for screening technologies for use in a particular application. Various other oxidative processes have been used to destroy free cyanide. Oxidants that have been employed in those processes include potassium permanganate, air, and sulfur dioxide [6]. All these processes have been implemented on a full-scale basis. Oxygen has also been successfully used to oxidize free cyanide in laboratory bench-scale experiments [31]. Permanganate is a powerful oxidant for free cyanide, but chemical costs for a full scale application might be cost prohibitive. Air might be useful as an oxidant at elevated temperature and pressure in order to decompose cyanide at appreciable rates. Using free oxygen, Bernardin [31] oxidized free cyanide to cyanate, and subsequently to ammonia and carbon dioxide in the laboratory using a catalytic column of copper and activated carbon. Free cyanide reduction of 99% was achieved from an influent cyanide concentration of 100 mg/l. The presence of organics and strong metal–cyanide complexes, however, were shown to reduce the process efficiency through competitive oxygen demand, preferential adsorption, and column fouling. Chlorine dioxide gas has also been successfully used to oxidize free cyanide to nondetectable levels after stripping cyanide from solution using air sparged hydrocyclone (ASH) technology [32]. Both bench- and pilot-scale applications of chlorine dioxide in ASH have been proven effective and potentially economical for the destruction of free cyanide in solution and slurries. Finally, a chemical reduction approach for treatment of free cyanide has been tested as an alternative to chemical oxidation. Formaldehyde (CH2 O) has been demonstrated to react rapidly with free cyanide and reduce it to form nontoxic, biodegradable glyconitrile [33,34].

20.6 SUMMARY AND CONCLUSIONS • Free and weak metal–cyanide complexes can be destroyed using conventional oxidation technologies, which include alkaline chlorination, ozonation, and hydrogen peroxide treatment.

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X

X

X

Alkaline chlorination

Hydrogen peroxide

Ozonation

Technology

Free CN

X

X

X

WAD CN FeCN

Chemical applicability General description

This technology involves the oxidation and destruction of free and WAD forms of cyanide under alkaline pH (9–11) conditions. Cyanide (CN− ) oxidation to cyanate (CNO− ) occurs in 10–15 min in the presence of excess of ozone under alkaline conditions. The use of UV light to enhance radical formation and the presence of copper catalyst have each been shown to increase the rate of oxidation, and to further oxidize cyanate to CO2 and N2 at longer retention times

Hydrogen peroxide oxidation of free and WAD CN is effective under alkaline conditions, at elevated temperatures, and in the presence of a metal catalyst (Cu, Fe, Al, Ni) or formaldehyde. The patented Kastone process utilizes H2 O2 and formaldehyde to oxidize cyanide (CN− ) to cyanate (CNO− ) at 49 to 54◦ C/pH 10–12

This technology involves oxidation and destruction of free and WAD CN under alkaline pH (10.5 to 11.5) conditions. The chlorine is supplied either in liquid form or as solid NaClO or CaOCl2 , which could be generated on-site electrolytically. This technology is the oldest and most widely recognized cyanide destruction process based upon operational experience and engineering expertise

TABLE 20.7 Oxidation Technology Screening Matrix

$5–7/kg CN destroyed

$11/kg CN treated for a 4800 gpm system

$300K for a 500 gpm system

$1M for a 4800 gpm system

$875K for a 500 gpm system

WAD CN <1 mg/l and free CN <0.2 mg/l

1–10 mg/l total CN and <0.5 mg/l WAD CN for a total CN influent of 110–300 mg/l

<0.1 mg/l

$2/kg CN destroyed

O&M

Costs Capital

Achieveble treatment levels

Minimal

Minimal

Minimal

Waste mgmt.

Establisbed. Chemical feed and monitoring equipment commercially available

Established. Peroxidation Systems manufactures modular systems

Established. Chemical feed and monitoring equipment commercially available

Technology status

Ambient Temperature Oxidation Technologies 409

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X

X

X

X

Photocatalytic oxidation

SO2 /air oxidation

Technology

WAD CN

X

X

FeCN

Chemical applicability

Free CN

TABLE 20.7 Continued

This patented technology by INCO uses Zn, Ni, and Cd to precipitate FeCN, followed by oxidation of free and WAD CN using SO2 and air in the presence of copper catalyst. Acid produced in the SO2 /Air oxidation rxn. is neutralized with CaO at pH 7 to 10. For WAD CN, the following conditions are recommended: pH ∼9; CN− /Cu2+ mass ratio of 5:1; and CN− /SO2 mass ratio between 1:3 and 1:7

This technology involves the photodissociation of FeCN complexes and certain other metal–cyanide complexes in the presence of UV light. The liberated free CN from the photolysis rxn. is destroyed by chemical oxidation to CO2 and NO− 3 using either ozone or H2 O2 in the presence of TiO2 catalyst

General description $1.4M (UV-H2 O2 ) and $1.83M (UV-ozone) for a 25 gpm GW system

$210K for a 1 kilo ton tailings/day system

<0.5 mg/l CN for a CN influent >350 mg/l

Capital

<0.5 mg/l CN in 2 h rxn. time for a SPL leachate of 74 mg/l CN

Achieveble treatment levels

$1.36/ton of tailings treated

$0.28M/yr (UV-H2 O2 ) and $0.26M/yr (UV-ozone) for a 25 gpm GW system

O&M

Costs

Not available

∼$100K/yr for off-site transport and nonhazardous landfill disposal for 25 gpm system

Waste mgmt.

More than 40 licenses sold for full-scale INCO CN destruction technology to date

Limited field-scale implementation. Only 2 to 3 actual field applications documented

Technology status

410 Cyanide in Water and Soil

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• Cyanate, CNO− , is the primary product of oxidation. Further oxidation of cyanate to carbon dioxide requires longer reaction times and addition of excess oxidant. • Alkaline chlorination is the most widely used ambient temperature oxidation technology. There is substantial full-scale experience, especially in the electroplating and gold mining industries. • Higher pH (9.5 to 12) is required with the conventional oxidation technologies for fast reactions and to prevent generation of toxic CNCl or HCN gas. • Alkaline chlorination, ozonation, and peroxide oxidation technologies are well established, moderately expensive, and usually uncomplicated to implement in the field. • The most feasible approach for destroying strong metal–cyanide complexes such as iron– and cobalt–cyanide under ambient temperature and pressure conditions is by photocatalytic oxidation. • The presence of metals and metal–cyanide complexes in the waste stream will result in the formation of metal hydroxide sludges, which usually require additional management and treatment prior to disposal.

REFERENCES 1. Chamberlin, N.S. and Snyder, H.B., Technology of treating plating wastes, in Proceedings of the 10th Purdue Industrial Waste Conference, Purdue University, West Lafayette, IN, 1955, p. 277. 2. Clark, D.P., Poulter, L.W., Wilson, O.W., and Christensen, W.N., The treatment and analysis of cyanide wastewater, prepared for Air Force Engineering Center, Report No. AFCEC-TR-74-5, Thiokol Corporation, Tyndall AFB, FL, 1975. 3. Gurol, M.D. and Bremen, W.H., Kinetics and mechanism of ozonation of free cyanide species in water, Environ. Sci. Technol., 19, 804, 1985. 4. Streeben, L.L., Schornick, H.M., and Wachinski, A.M., Ozone oxidation of concentrated cyanide wastewater from electroplating operations, in Proceedings of the 35th Purdue Industrial Waste Conference, Purdue University, West Lafayette, IN, 1980, p. 655. 5. Theis, T.L., Young, T.C., Schaefer, R.J., and Tudman, S., Advanced oxidation of iron cyanides, in Proceedings of WEFTEC 97, Vol. 3 Symposium on Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1997, p. 135. 6. Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Surprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Corp, N.D., Ed., Noyes Data Corp., Park Ridge, NJ, 1988. 7. Shelton, S.P., Examination of treatment methods for cyanide wastes, Report No. NADC-78198-60, Naval Material Command, Washington, DC, 1979. 8. Hartinger, L., Handbook of Effluent Treatment and Recycling for the Metal Finishing Industry, 2nd ed., Finishing Publications, Herts, U.K., 1994. 9. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. 10. Hassan, S.Q., Vitello, M.P., Kupferle, M.J., and Grosse, D.W., Treatment technology evaluation for aqueous metal and cyanide-bearing hazardous waste, J. Air Waste Manage. Assoc., 41, 710, 1991. 11. Wedl, D.J. and Dfaulk, R.J., Cyanide destruction in plating sludges by hot alkaline chlorination, Metal Finish., 89, 33, 1991. 12. Patterson, J.W., Cyanide, in Industrial Wastewater Treatment Technology, 2nd ed., ButterworthHeinemann, Boston, MA, 1985, p. 115. 13. Altmayer, F., Improving the operation of cyanide destruction systems, Plating Surf. Finish, 75, April 20, 1988. 14. Herlacher, M.F. and McGregor, F.R., Photozone destruction of cyanide waste at Tinker AFB (pilot plant results), Paper No. 870746, in Proceedings of 23rd Annual Aerospace/Airline Plating and Metal Finishing Forum and Exposition, Jacksonville, FL, 1987. 15. Gurol, M.D. and Holden, T.E., The effect of copper and iron complexation on removal of cyanide by ozone, Ind. Eng. Chem. Res., 27, 1157, 1988.

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16. Layne, M.E., Singer, P.C., and Lidwin, M.I., Ozonation of thiocyanate, in Proceedings of Conference on Cyanide and the Environment, Tucson, AZ, 1984, p. 433. 17. USEPA, Managing cyanide in metal finishing, Capsule Report, EPA 625/R-99/009, U.S. Environmental Protection Agency, Office of Research and Development, Cincinnati, OH, 2000. 18. Botz, M., Devuyst, E.A., Mudder, T., Norcross, R., Ou, B., Richins, R., Robbins, G., Smith, A., Steiner, N., Stevenson, J., Waterland, R., Wilder, A., and Zaidi, A., An overview of cyanide treatment and recovery methods, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd., London, 1998. 19. Evans, F.L., Ozone in Water and Wastewater Treatment, Ann Arbor Science, Ann Arbor, MI, 1972. 20. Asperger, S., Kinetics of the decomposition of potassium ferrocyanide in ultra violet light, Trans. Faraday Soc., 48, 617, 1952. 21. Gaspar, V. and Beck, M.T., Kinetics of the photoaquation of hexacyanoferrate (II) ion, Polyhedron, 2, 387, 1983. 22. Schaefer, R.J., Photocatalytic treatment of cyanide in aluminum potlining leachate using ozone as an oxidizing agent, M.S. thesis, Clarkson University, Potsdam, NY, 1996. 23. Scott Rader, W., Solujic, L., Milosavljevic, E.B., and Hendrix, J.L., Sunlight-induced photochemistry of aqueous solutions of hexacyanoferrate-(II) and -(III) ions, Environ. Sci. Technol., 27, 1875, 1993. 24. Zhao, J., The treatment of cyanide-bearing wastes at manufactured gas plants, M.S. thesis, Clarkson University, Potsdam, NY, 1994. 25. Knutsen, K.C., Leaching behavior and treatment of cyanide-bearing wastes at manufactured gas plants, M.S. thesis, Clarkson University, Potsdam, NY, 1992. 26. Devuyst, E.A., Robbins, G., Vergunst, R., Tandi, B., and Iamarino, P.F., INCO’s cyanide removal technology working well, Mining Eng., Feb., 205, 1991. 27. Devuyst, E.A., Tandi, B., and Conard, B.R. Treatment of cyanide–ferrocyanide effluents, U.S. Patent No. 4,615,873, 1986. 28. Scott, J. and Ingles, J., State of the art processes for the treatment of gold mill effluents, Mining, Mineral and Metallurgical Processes Division, Environment Canada, Ontario, Canada, 1987. 29. Devuyst, E.A., Vergunst, R.D., Iamarino, P.F., and Agius, R.J., Recent applications of the INCO SO2 /air cyanide removal process, in Proceedings of the Conference of 94th Annual General Meeting of the CIM, Montreal, CA, 1992. 30. Mudder, T., Editorial comment: minerva, Mining Environ. Manage., 9, 3, 2001. 31. Bernardin, F.E., Cyanide detoxification using adsorption and catalytic oxidation on granular activated carbon, J. Water Pollut. Control Fed., 45, 221, 1973. 32. Pargar, J.R. and Miller, J.D., Cyanide recovery/destruction using air sparged hydrocyclone technology, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, New Orleans, LA, 2001, p. 363. 33. Stone, D.E., Reduction of weak acid dissociable cyanide using formaldehyde, Iron Steel Engineer, 75, 51, 1998. 34. Colin, F., d’Ambrosio, G., and Grapin, F., Specific removal of cyanides in steelwork effluents, Cahiers Inf. Tech. — Rev. Metal., 88, 979, 1991.

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Technologies 21 Separation for Treatment of Cyanide David A. Dzombak, Rajat S. Ghosh, George M. Wong-Chong, and John R. Smith CONTENTS 21.1

Adsorption Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.1.1 Activated Carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.1.2 Synthetic Ion Exchange Resin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.1.3 Activated Alumina . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2 Precipitation Technologies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2.1 Iron–Cyanide Precipitation Reactions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2.2 Iron–Cyanide Precipitation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2.3 Operation and Regulatory Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.3 Air Stripping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.3.1 Counter-Current Air Stripping. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.3.2 Recovery and Reuse of Cyanide in Hydrometallurgical Gold Mining . . . . . . . . . 21.4 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

413 414 418 421 423 423 424 428 429 429 430 434 435

While low-temperature and thermal oxidation technologies are employed most commonly for destruction of cyanide species in water, soils, sludges, and solids (see Chapters 20 and 22), physical– chemical separation of cyanide species is an alternative, sometimes less expensive, treatment approach for cyanide-bearing waters. This is especially the case for treatment of strong metal– cyanide species, which often are not as readily oxidized as free cyanide and weak-acid-dissociable (WAD) cyanide. The reactivity of cyanide species on certain kinds of solid surfaces, activated carbon in particular, is exploited for adsorptive removal from water. In addition, the ability of cyanide to react with certain metals, especially iron, to form relatively insoluble solids is exploited for precipitative removal of free and metal complexed cyanide. Finally, air stripping can be used to remove free cyanide from water, as HCN is a volatile species. In this chapter, we examine adsorption, precipitation, and air stripping treatment approaches that have been used to treat various cyanide-bearing waters.

21.1 ADSORPTION TECHNOLOGIES Adsorption of cyanide species from water onto solid phases has not been extensively studied, but it is known that free cyanide and metal–cyanide species exhibit surface reactivity with certain mineral solids and with activated carbon. From the limited studies that have been performed, it appears that free cyanide (HCN, CN− ) has little or no tendency to adsorb on inorganic minerals such as iron oxides [1] and sand [2], but does adsorb on activated carbon [3] as well as on soil organic carbon [4,5]. 413

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Adsorption on activated carbon has been used to recover free cyanide and gold–cyanide complexes from heap leach solutions in hydrometallurgical gold mining [6,7]. Free cyanide has also been removed from solution with synthetic ion exchangers, but usually after first complexing the cyanide with iron [3]. Metal–cyanide species, on the other hand, are known to adsorb on inorganic minerals, activated carbon and soil organic carbon, and on ion exchange resins. The adsorption of ferro- and ferricyanide species, the most common metal–cyanide species encountered in water and soil systems, has received the greatest attention. In this section, we focus on adsorbents that have been used at some scale for treating cyanidecontaminated waters. The adsorbents discussed are activated carbon, synthetic anion exchange resin, and granular activated alumina. Applications of these adsorbents to remove free cyanide and iron– cyanide species from simple and complex cyanide-bearing waters are examined.

21.1.1 ACTIVATED CARBON The ability of granular and powdered activated carbon to remove cyanide from water has been exploited in industrial wastewater treatment [8–11]. Activated carbon has been shown capable of adsorbing free cyanide, especially as HCN [3,10,12], and also metal cyanides, especially iron cyanides [13,14]. A significant factor motivating studies of cyanide adsorption by activated carbon was the development of a process involving use of activated carbon as a medium and catalyst for oxidative destruction of cyanide [12,15,16]. The oxidation technology employing activated carbon is not covered here (see Chapter 20), but knowledge gained about activated carbon as a cyanide adsorbent in the development and application of this technology is discussed. Most testing and application of activated carbon for removal of cyanide has been done with granular activated carbon (GAC), for use in fixed-bed reactors. Many different types of GAC are available. These porous, high surface area materials have predominantly non-polar carbon surfaces but can be treated in various ways for installation of desired kinds of reactive surface functional groups. For example, activated carbons are often “impregnated” with reactive metals, or reacted with oxygen to provide an oxidized carbon surface. Equilibrium adsorption testing with GAC materials and free cyanide has indicated that the extent of adsorption is not dependent on pH [3,10], at least for pH < 10 where free cyanide is predominantly in the neutral HCN form [10], and that the adsorption capacity of the GAC is in the range of 2 to 5 mg CN/g GAC for the aqueous cyanide concentration range of typical interest, that is, less than 100 mg/l [3,12,15]. This is illustrated in Figure 21.1, which presents an isotherm for adsorption of free cyanide on a commonly used GAC from Calgon Carbon. As seen there, for aqueous free cyanide concentrations up to 1 mg/l, cyanide adsorption follows a Freundlich isotherm and the maximum adsorbed concentration is about 2.2 mg CN/g GAC. The presence of other solutes that can form complexes with cyanide in solution or compete for adsorption sites influences cyanide uptake by GAC. Figure 21.2 presents the results of adsorption experiments performed with solutions containing free cyanide alone, as well as solutions with Al3+ , Fe3+ , Ca2+ , and Mg2+ salts added. It is clear that the presence of relatively large concentrations of these reactive cations reduces free cyanide adsorption on GAC. Ferric iron is the only one of the ions tested known to form complexes with cyanide (see Chapter 5), which would serve to keep cyanide in solution. Precipitation of the solid ferric hydroxide also occurred in the experiments with Fe3+ and likely blocked some of the GAC surface area. The same mechanism explains some of the inhibition of adsorption in the Al3+ experiments, in which aluminum hydroxide formed. The mechanism for the inhibition of cyanide adsorption induced by Ca2+ and Mg2+ is unclear, however. The presence of organic compounds, which generally have more affinity for GAC than inorganic ions or compounds, can also compete for adsorption sites, and thus inhibit cyanide adsorption. This may be seen in Figure 21.3, which shows the effect of humic acid on cyanide adsorption to GAC. The presence of some cyanide-complexing metals enhances cyanide adsorption on GAC considerably. Bernardin [12] demonstrated that additions of copper, cobalt, iron, nickel, and zinc to solutions

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Log [cyanide uptake (mg CN/g GAC)]

0.4

0.3

0.2

0.1 –2

–1.5

–1

–0.5

0

Log (Equilibrium concentration of CN, mg/l)

FIGURE 21.1 Freundlich isotherm for adsorption of free cyanide on Calgon Carbon GAC FS-400. GAC dose range = 100 to 600 mg/l. Total CN = 1.0 mg/l. pH = 9. T = 22◦ C. I = 0.01 M. Reaction time = 33.5 h. (Source: Guo, R. et al., Water Environ. Res., 65, 640, 1993. With permission.)

Percentage of total cyanide adsorbed (%)

40

30

20

No foreign ions added 10 mg/l AI3+ 10 mg/l Fe3+

10

100 mg/l Ca2+ 100 mg/l Mg2+

0 4

5

6

7

8

9

pH

FIGURE 21.2 Effect of Al3+ , Fe3+ , Ca2+ , and Mg2+ ions on free cyanide adsorption by Calgon Carbon GAC FS-400 as a function of pH. GAC dose = 600 mg/l. Total CN = 1.0 mg/l. T = 22◦ C. I = 0.01 M. Reaction time = 2.0 h. (Source: Guo, R. et al., Water Environ. Res., 65, 640, 1993. With permission.)

of free cyanide significantly increased adsorption of cyanide. For example, in adsorption testing with GAC (Calgon Carbon Filtrasorb 300 and Filtrasorb 400) and various cyanide bearing wastes, the adsorption capacity of the GAC was determined to be in the range of 2 to 3 mg CN/g GAC. When cupric ion (Cu2+ ) was added to the solutions at a molar Cu:CN ratio of approximately 1:1, the adsorption capacity of the GAC for cyanide was increased to 25 mg CN/g GAC. This was attributed to the formation of copper-cyanide complexes, and the adsorption of the copper on the GAC along with the bound cyanide. The important role of the copper was confirmed in adsorption experiments with the copper alone and cyanide alone. For equilibrium solution concentrations of 20 mg/l in each

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Percentage of total cyanide adsorbed (%)

40

30

20

10 10 mg/l NTA 10 mg/l humic acid 0 0

2

4 6 8 Concentration of complexing agent (mg/l)

10

FIGURE 21.3 Effect of humic acid on free cyanide adsorption by Calgon Carbon GAC FS-400 as a function of humic acid concentration. GAC dose = 600 mg/l. Total CN = 1.0 mg/l. pH = 9. T = 22◦ C. I = 0.01 M. Reaction time = 2.0 h. (Source: Guo, R. et al., Water Environ. Res., 65, 640, 1993. With permission.)

case, the adsorbed concentration was 5.8 mg/g for Cu2+ , and 1.0 mg/g for CN− . Similar results were found for Co2+ , Ni2+ , and Zn2+ . Ferric iron (Fe3+ ) increased cyanide adsorption on GAC, but to a lesser extent than the other metals tested. The ability of metals, especially copper, to enhance the adsorption of cyanide on activated carbon has been investigated and exploited by a number of investigators [10,11,15]. Copper, either impregnated in GAC or added to contaminated solutions to be treated, has been used to enhance adsorption of cyanide in a catalytic oxidation process in which cyanide adsorbed on the carbon is destroyed [8,11,12,15,16]. The adsorption of metal-cyanides on granular activated carbon is dependent on pH. Chank [17] studied the adsorption of ferrocyanide, Fe(CN)4− 6 , also known as hexacyanoferrate, on activated carbon as a function of pH. Results are shown in Figure 21.4, where it may be seen that adsorption of ferrocyanide was greatest at low pH, and decreased as pH increased. Above pH 10, ferrocyanide adsorption was negligible. In hydrometallurgical gold mining (heap leaching), activated carbon is used for recovery of gold cyanide through direct contact with cyanide-leached ore slurry (“pulp”) in countercurrent flow. This is known as the carbon-in-pulp (CIP) or carbon-in-leachate (CIL) process, depending on the location in which the activated carbon is added in the leach circuit [6,7,18]. The activated carbon is recovered from the leach slurry by screening. Gold cyanide adsorbed on the activated carbon is eluted with a solution of cyanide and sodium hydroxide. Mining operations that utilize carbon adsorption typically have the full range of on-site processing equipment for carbon systems (e.g., loading, storing, conveying, and regeneration). Carbon adsorption can be applied using conventional dual carbon columns operating in a lead-lag mode, or in a traveling bridge adsorption system. Figure 21.5 and Figure 21.6 present schematic flow diagrams of these process configurations. The traveling bridge system includes automated backwashing of the filter and carbon layers. It is most economical for processing large volume process flows (e.g., an inflow of 20,830 m3 /day at the Homestake Mine, Lead, SD), whereas the dual-carbon-column configuration is best suited to smaller flows. Botz and Mudder [19] present summary data on six mining operations for which carbon adsorption treatment was evaluated or commercially applied to the processing of mining process waters. Table 21.1 summarizes the performance data reported for several

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Percent hexacyanoferrate(II) removed

100 90 80 70 60 50 40

I = 0.01 M KNO3

30

10 mg/l hexacyanoferrate(II)

20

20 mg/l hexacyanoferrate(II) 30 mg/l hexacyanoferrate(II)

10

40 mg/l hexacyanoferrate(II)

0 0

2

4

6 pH

8

10

12

FIGURE 21.4 Percent hexacyanoferrate removed by granular activated carbon vs. pH. Hexacyanoferrate (10 to 40 mg/l) in 0.01 MKNO3 solution. Granular activated carbon = 10 g/l. (Source: Chank, J.K., M.S. thesis, Clarkson University, 1997.)

GAC Adsorber 1

GAC Adsorber 2

Influent

FIGURE 21.5 Schematic diagram of two-column, lead-lag carbon adsorber system.

of these operations [19]. These data show that carbon adsorption was effective in removing free and WAD cyanide. Some heavy metal (e.g., copper and zinc) removal was also observed. Based on industry experience, Botz and Mudder [19] provide specific design recommendations for carbon adsorption treatment of gold mine process waters. While activated carbon treatment of cyanide-bearing waters has been demonstrated at bench, pilot, and full-scale, it is not widely practiced for wastewater treatment beyond the gold mining industry. Fixed-bed GAC treatment is relatively expensive. For example, Botz and Mudder [19] reported that for treatment of gold mining industry process water flows up to 500 gpm, the constructed plant cost (including carbon columns, heated structure, and piping) was approximately $30,000 (1997 cost basis) for each 100 gpm to be treated. Operating costs are also significant, for periodic carbon replacement (carbon cost is $0.80 to $0.90 per pound [20]) and for additives to prevent column

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Traveling bridge

Washwater hood Influent

Filter with underdrain

GAC Effluent

FIGURE 21.6 Schematic diagram of traveling bridge carbon adsorption system.

TABLE 21.1 Performance of Activated Carbon Systems Treating Gold Mining Process Watersa Process water

Influent CNT (mg/l)

Effluent CNT (mg/l)

Influent WAD CN (mg/l)

Effluent WAD CN (mg/l)

0.98

0.20

0.70b

0.02b

NAc

NA

1.08

0.10

NA

NA

5.53

0.06

Biotreated tailings water: pilot tests H2 O2 treated rinse solution: pilot tests H2 O2 treated rinse solution: full-scale

a Source: Data from Botz, M. and Mudder, T., Cyanide Monograph, Mudder, T., Ed., Mining Journal Books,

Ltd, London, 1998. b Reported as free cyanide. c NA: no data available.

plugging. Carbon replacement rather than regeneration is often required because cyanide-bearing waters often contain metals that make carbon regeneration uneconomical [9]. Further, implementation of fixed-bed GAC treatment often requires filtration pre-treatment in order to remove suspended solids that can clog the bed, which adds to the cost of the system. The use of powdered activated carbon in complete-mix reactors (such as activated sludge reactors) for adsorption of cyanide can be effective for waters with low organic carbon concentrations [11], though sludge disposal may be expensive and complicated from a regulatory standpoint due to the presence of cyanide and metals.

21.1.2 SYNTHETIC ION EXCHANGE RESIN Ion exchange is not used commonly for treatment of cyanide-bearing wastewaters [8], but it has been evaluated and used for removal of metal–cyanide complexes from solution [8,9]. The most common applications are use of strong base anion exchangers for removal of iron–cyanide (Fe(CN)3− 6 − and Fe(CN)4− ) and gold–cyanide (Au(CN) ) complexes from solution. Ion exchange has been 2 6 investigated extensively for use in recovery of gold–cyanide complexes in process water from hydrometallurgical gold mining [6,21–23]. To the extent that ion exchange is used to treat waters with free

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cyanide, the free cyanide is typically first complexed with iron or another metal and then contacted with an anion exchange resin [6,9,24]. Synthetic ion exchanger beads consist of a cross-linked polymer matrix to which, charged functional groups are attached by covalent bonding [25]. The usual matrix is polystyrene cross-linked with 3 to 8% divinylbenzene for structural stability [25]. Negatively charged acid groups are attached for cation exchange; positively charged base groups are attached for anion exchange. A common type of functional group on anion exchangers, for example, is the strongly basic quaternary amine group (RN(CH3 )+ 3 ), where R represents a hydrocarbon group in the resin matrix [9,25]. Ion exchange involves the exchange of one bound ion for another at the charged sites on the exchanger surface. Consider, for example, an anion exchanger with quaternary amine groups as exchange sites, where all sites are initially occupied with chloride (Cl− )ions. The exchange of ferrocyanide for chloride would occur as follows [9]: 4− + 4− − − 4[RN(CH3 )+ 3 ] − Cl + Fe(CN)6 = [RN(CH3 )3 ]4 − Fe(CN)6 + 4Cl

(21.1)

Exchangers can be provided with different initial counterions. If a quaternary amine group anion exchanger is in the “sulfate form” initially rather than the “chloride form,” then one sulfate ion will 2− counteract the charge of two exchange sites: [RN(CH3 )+ 3 ]2 − SO4 . Gold–cyanide complexes are removed from gold mining process waters via reactions on such exchange sites [6]: 2− − + − 2− [RN(CH3 )+ 3 ]2 − SO4 + 2Au(CN)2 = 2[RN(CH3 )3 ] − Au(CN)2 + SO4

(21.2)

These exchange reactions, for removal of metal cyanides from solution, are sometimes referred to as the “loading step” in the ion exchange process. When the fraction of ion exchange sites occupied with the target ion becomes large, and the ion exchange capacity is nearing exhaustion, the ion exchanger is usually regenerated by contacting it with a small volume of concentrated solution containing an ion capable of displacing the target ion from the exchanger. For example, the quaternary amine group anion exchanger loaded with ferrocyanide can be regenerated with aqueous sodium chloride as follows [9]: 4− + 4− − + [RN(CH3 )+ 3 ]4 − Fe(CN)6 + 4NaCl = 4[RN(CH3 )3 ] − Cl + Fe(CN)6 + 4Na

(21.3)

Regeneration of the same anion exchanger loaded with gold cyanide via the reaction in Equation (21.2) is usually performed in two steps, with the first “stripping” step designed to isolate the gold cyanide for recovery, and the second to regenerate the ion exchanger [6]. The stripping step is conducted using a metal cyanide compound with similar exchange properties as the gold cyanide: − 2− + 2− − 2[RN(CH3 )+ 3 ] − Au(CN)2 + Zn(CN)4 = [RN(CH3 )3 ]2 − Zn(CN)4 + 2Au(CN)2

(21.4)

Sulfuric acid is then used to regenerate the exchanger into the sulfate form: 2− + 2− 2+ [RN(CH3 )+ + SO2− 3 ]2 − Zn(CN)4 + 2H2 SO4 = [RN(CH3 )3 ]2 − SO4 + Zn 4 + 4HCN (21.5)

The formation of HCN is common in acid regeneration of ion exchangers with adsorbed metal– cyanide species [8]. This is a potential problem in that HCN is volatile and highly toxic. Ion exchange treatment of water is usually implemented using fixed-bed reactors. Water pretreated for removal of particles and organics is passed through a bed of the ion exchange material. When the ion exchange capacity of the bed nears exhaustion, it is regenerated. In hydrometallurgical gold mining, anion exchange resins are being investigated as an alternative to activated carbon for recovery of gold cyanide from cyanide-leached ore slurry [18]. This is known

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TABLE 21.2 Results of Anion Exchange Treatment of Electroplating Wastewater Pretreated with FeSO4 to Convert Free Cyanide to Iron–Cyanide Complexesa CNT (mg/l)

NATCb CN (mg/l)

Cd (mg/l)

Cu (mg/l)

Ni (mg/l)

Zn (mg/l)

Fe (mg/l)

Raw waste Anion exch. eff.

350 25.0

33.0 NAc

29.5 1.41

68.7 0.30

14.8 0.00

12.2 0.05

13.3 0.50

Fe hydrox. sludge

9310

NA

(mg/kg wet sludge) 48.1 8.70

68.7

1800

38,500

Sample

a Source: Adapted from Hassan, S.Q., Vitello, M.P., Kupferle, M.J., and Grosse, D.W., J. Air Waste Manage.

Assoc., 41, 710, 19. b NATC = not amenable to chlorination. c NA = not analyzed.

as the resin-in-pulp (RIP) process. The process is not commonly used yet, but interest in it is increasing because of the promise of increased recoveries [18] and because of expanded interest in recycling the cyanide extractant [21]. There is full-scale experience with use of ion exchange to remove metal–cyanide complexes to <1 mg/l from gold mining and other industrial wastewaters containing total cyanide concentrations ranging from tens to hundreds of milligrams per liter [9,22,26]. For an ion exchanger with selectivity for cyanide or a metal–cyanide species in a particular water, the effluent concentration achievable is a matter of design. Higher resin-to-water ratios and contact time can yield lower treated effluent concentrations [25]. Water composition will determine the degree of selectivity. An important parameter governing selectivity of resins for metal cyanides is pH. Moore [27], for example, demonstrated that the removal of cyanide and zinc cyanide from electroplating wastewater by anion exchange decreased significantly at pH > 10. Hassan et al. [24] studied at bench-scale the use of anion exchange to remove cyanide from electroplating wastewater. The treatment strategy involved addition of ferrous iron as FeSO4 in order to form iron–cyanide complexes in solution, and then to pass the water containing the iron– cyanide complexes through a column of anion exchange resin (Amberlite IRA-958) to remove the iron–cyanide complexes. Solution pH was controlled at pH 8.5 in the FeSO4 addition step. The cyanide mass loading rate to the ion exchange column was in the range 1.2 to 2.3 lb CN/ft3 of packed resin. Cyanide in electroplating wastewaters typically is mostly present as free cyanide or weak metal–cyanide complexes, and such was the case with the water employed in this work. As shown in Table 21.2, the electroplating wastewater contained 350 mg/l total cyanide, only a small fraction of which (33 mg/l) was strongly complexed and not amenable to destruction by chlorination. Total cyanide was removed in the anion exchange column from 350 to 25 mg/l (Table 21.2). It was found that upon addition of the ferrous sulfate, an iron hydroxide solid formed, which was settled and filtered prior to the anion exchange step. A significant amount of cyanide was removed with the iron hydroxide sludge (Table 21.2), through mechanisms not identified in the study. While the concentration of cyanide in the anion exchange effluent was relatively high, this presumably could have been improved by reducing the cyanide mass loading rate to the column, a design variable. USFilter Recovery Services, Inc. (Roseville, MN) markets a Purolite A-400 anion exchange resin for commercial removal of metal–cyanide complexes, including iron cyanide, from wastewaters. This resin is supplied in pre-packed tanks, ranging from 1 to 200 ft3 of bed volume, and exchanges metal–cyanide complex anion for chloride (Cl− ), sulfate (SO2− 4 ), and alkalinity (assumed to be bicarbonate). A pre-filter is required to filter the influent prior to passing through the ion-exchange

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FIGURE 21.7 USFilter Corp.

421

Schematic diagram for ion exchange system for metal-cyanide removal. Courtesy of

columns in order to eliminate any interference between total suspended solids (TSS) and the resin functional groups. Figure 21.7 depicts the ion exchange system for cyanide removal. A great deal of research and fieldwork has been conducted on the regeneration of anion exchange materials after their use for removal of metal cyanides from solution. Regeneration and reuse of ion exchange material is critical to economic implementation of the process. Work on regeneration and reuse has included investigation of eluents, efficiency of recovery, and loss of ion exchange capacity with multiple cycles of regeneration. Commonly used eluents are concentrated solutions of sodium chloride, zinc cyanide, ammonium and potassium thiocyanate, sodium hydroxide, and thiourea [22,23,26]. The effectiveness of these eluents in removing cyanide compounds adsorbed on anion exchangers depends greatly on the solution conditions and exchanger properties. Recoveries are never complete, however, and sometimes can be very poor. In regeneration of an anion exchanger used to remove iron cyanide from various gold mill effluents, for example, Vachon [26] achieved recoveries as low as 40% in regeneration with a 15% NaCl solution. In repeated exchange-regeneration tests with a particular effluent, Vachon [26] further observed that the resin lost 25% of its exchange capacity in the first cycle, and the additional loss in exchange capacity for subsequent cycles was approximately 1% per cycle. The loss of exchange capacity upon regeneration is reported to be a common problem which limits the use of strong base anion exchangers [9]. Another factor limiting use of ion exchange for removal of cyanide species from water is concern about formation of highly toxic HCN gas upon resin regeneration [8]. Thus, ion exchange is a technology that has been demonstrated to be capable of removing cyanides, especially metal–cyanide compounds, from aqueous solution. The technology has not been widely adopted for this purpose, however, because of the relatively high cost of ion exchange materials which is approximately $2.00/lb, or more than twice the cost of activated carbon [20,28]; the inability to regenerate the material after use with a high degree of efficiency; and concern about HCN gas formation during resin regeneration.

21.1.3 ACTIVATED ALUMINA Activated alumina, a granular adsorbent used in water and wastewater treatment for removal of ionic chemical species from water [25], has been shown to adsorb ferrocyanide [17,29]. The hydroxyl groups on the surface of this aluminum oxide, designated ≡AlOH, are the reactive entities. Adsorption of ferrocyanide, Fe(CN)4− 6 , on the activated alumina surface occurs by surface complexation, for example, + 4− ≡AlOHo + H+ + Fe(CN)4− 6 = ≡AlOH2 − Fe(CN)6

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4− where ≡AlOHo and ≡AlOH+ 2 − Fe(CN)6 are surface complexes. Equation (21.6) was proposed by Chank [17]. Other metal oxides, such as the iron oxide goethite, also have reactive surface hydroxyl groups and the ability to adsorb metal–cyanide species like ferrocyanide [1]. Adsorption of ions on aluminum oxides and other metal oxides is strongly pH dependent [30]. Surface hydroxyl groups on activated alumina can bond with and release H+ ions, and thus are amphoteric, having both acid and base properties.

≡AlOHo + H+ = ≡AlOH+ 2 −

≡AlOH = ≡AlO + H o

(21.7)

+

(21.8)

The H+ exchange at the surface alters the surface charge, which in turn affects the energy of adsorption for ions. Changes in pH also affect the solution speciation, including the adsorbing species, which further alters the surface reactions. These factors are responsible for the strong pH dependence of ion adsorption on oxides. Adsorption of cations usually increases with pH and is maximum at higher pH values, while adsorption of anions is greatest at low pH and decreases with pH [30]. The strong dependence on pH is typically illustrated by plotting adsorbed ion concentrations versus pH. Data from Chank [17] for adsorption of ferrocyanide vs. pH are presented in Figure 21.8. The “pH adsorption edge” plots shown in Figure 21.8 illustrate that granular activated alumina has a high affinity for ferrocyanide at pH < 8. Complete removal of ferrocyanide from solution was achieved for ferrocyanide concentrations in the range of 10 to 40 mg/l, and for an alumina concentration of 10 g/l. Similar experiments carried out with granular activated carbon, shown in Figure 21.4, revealed that activated alumina has a greater adsorption capacity for ferrocyanide than activated carbon, though both solids exhibit significant capability to adsorb ferrocyanide at lower pH values. Despite the significant adsorption capacity of activated alumina for metal cyanides, and the availability of the technology at full scale, it has seen little use for treatment of cyanide-bearing waters. Activated carbon and ion exchange technologies are more commercially accessible and are the adsorption technologies investigated first when there is interest in use of adsorption. Considering the favorable adsorptive properties of granular activated alumina, and its competitive cost, especially

Percent hexacyanoferrate(II) removed

100 90 80 70 60 50 I = 0.01 M KNO3

40

10 mg/l hexacyanoferrate(II) 20 mg/l hexacyanoferrate(II) 30 mg/l hexacyanoferrate(II) 40 mg/l hexacyanoferrate(II)

30 20 10 0

0

2

4

6 pH

8

10

12

FIGURE 21.8 Percent hexacyanoferrate removed by granular activated alumina vs. pH. Hexacyanoferrate (10 to 40 mg/l) in 0.01MKNO3 solution. Granular activated alumina = 10 g/l. (Source: Chank, J.K., M.S. thesis, Clarkson University, 1997.)

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relative to ion exchange [25], it warrants more consideration for treatment of cyanide contamination in water.

21.2 PRECIPITATION TECHNOLOGIES Cyanide reacts with many metals to form solution complexes and a wide range of solids (Chapter 5). The low solubility of some of these solids, especially the iron–cyanide solids, has been exploited for removal of cyanide from solution by precipitation. As is the usual case for application of precipitation technology, the process is most economical for treatment of waters with fairly high concentrations of cyanide, for example, greater than 50 mg/l total cyanide. Precipitation has been employed primarily for removal of metal cyanides, especially strong metal–cyanide complexes like iron cyanides, since these cyanide species are difficult to oxidize. Precipitation of iron–cyanide solution species through formation of iron–cyanide solids is the most common application and hence the focus of this section.

21.2.1 IRON–CYANIDE PRECIPITATION REACTIONS The most common iron–cyanide reactions exploited for removal of cyanide species from solution are those leading to the formation of the pure iron cyanide solids Prussian Blue, Fe4 (Fe(CN)6 )3 (s) and Turnbull’s Blue, Fe3 (Fe(CN)6 )2 (s), which are blue in color and have different proportions of Fe2+ and Fe3+ [31]. Other iron cyanide solids can form with different ratios of Fe2+ and Fe3+ , including Prussian Brown, Berlin Green, and Berlin White, though these solids are unstable in most aqueous systems ([31]; and Chapter 5). The formation of Prussian Blue and Turnbull’s Blue can be represented by the following reactions: 4Fe3+ + 3Fe2+ + 18CN− = Fe4 (Fe(CN)6 )3 (s) 2Fe

3+

+ 3Fe

2+

−

+ 12CN = Fe3 (Fe(CN)6 )2 (s)

(21.9) (21.10)

Writing the reactions in terms of the combination of the constituent ions and molecules indicates the ratio of Fe2+ and Fe3+ in the product solids. The reactions can also be expressed, however, in terms of reactants that reflect more closely the actual combining species, such as ferrocyanide, Fe(CN)4− 6 , 3− or ferricyanide, Fe(CN)6 . 4Fe3+ + 3Fe(CN)4− 6 = Fe4 (Fe(CN)6 )3 (s)

(21.11)

3Fe2+ + 2Fe(CN)3− 6 = Fe3 (Fe(CN)6 )2 (s)

(21.12)

It may be seen from Equations (21.11) and (21.12) that iron–cyanide solid precipitation can be used for removal of dissolved iron–cyanide species from solution by adding additional iron to form ferric ferrocyanide solids. Numerous other metals can react with iron cyanide to form mixed-metal cyanides, including copper, cobalt, cesium, nickel, and zinc. The solid Zn2 Fe(CN)6 (s) is reported to form upon addition of Zn2+ to solutions of Fe(CN)4− 6 [32,33] 2Zn2+ + Fe(CN)4− 6 = Zn2 Fe(CN)6 (s)

(21.13)

Copper has been used in a similar manner to remove Fe(CN)4− 6 from solution [34–36]: 2Cu2+ + Fe(CN)4− 6 = Cu2 Fe(CN)6 (s)

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The removal of ferricycanide, Fe(CN)3− 6 , via precipitation with copper has also been reported [32]: 3Cu2+ + 2Fe(CN)3− 6 = Cu3 (Fe(CN)6 )2 (s)

(21.15)

Precipitation with ferrocyanide has been used for removal of cesium from solution, usually by coprecipitation with copper or zinc ferrocyanide, resulting in formation of a multi-metal solid [32,36]: 2Cs+ + Cu2+ + Fe(CN)4− 6 = Cs2 CuFe(CN)6 (s)

(21.16)

2Cs+ + Zn2+ + Fe(CN)4− 6

(21.17)

= Cs2 ZnFe(CN)6 (s)

The formation of various multi-metal cyanide solids involving cobalt, such as Co2 Fe(CN)6 (s) and Cs2 CoFe(CN)6 (s), has also been exploited for cobalt and cesium removal from solution [32].

21.2.2 IRON–CYANIDE PRECIPITATION PROCESSES Precipitation processes for water treatment can be conducted in batch or continuous processes. The required steps in the process are rapid mixing of the chemical precipitation agents, slow mixing for particle formation and aggregation (flocculation), and gravity sedimentation of the precipitates. For iron–cyanide precipitation, the chemical precipitation agents are acid or base to adjust the pH to the optimal pH range, and the salt of a metal ion to react with the dissolved iron–cyanide species and form the desired solid. A schematic representation of a continuous process for precipitation of Prussian Blue or Turnbull’s Blue is given in Figure 21.9. In this example process, acid or base is added to adjust to near-neutral pH, and then ferrous sulfate is added to provide excess ferrous iron to induce formation of iron-cyanide solids. As indicated in Chapter 5, the formation and solubility of iron–cyanide solids is strongly dependent on pH and pe. Formation of Prussian Blue and Turnbull’s Blue is favored at mid-range to low pH, with Prussian Blue formed under more oxidizing conditions (higher pe) and Turnbull’s Blue formed under more reducing conditions (lower pe). Stable iron–cyanide solids under equilibrium conditions have been studied [31], and such data provide guidance about optimal conditions for precipitation treatment processes. While knowledge of the equilibrium properties of the precipitating solids is useful for design of precipitation treatment processes, kinetics of reaction, the complexity of the aqueous solution, and other factors affect actual performance. Bench and pilot-scale testing with the water of interest is needed to develop a precipitation treatment system that provides the desired removal in the most cost-effective manner. This certainly is the case in most iron–cyanide treatment applications, for FeSO4

Influent with dissolved Fe–CN

Acid Effluent Mix tank

Mix/Floc tank

Settling tank

Sludge

FIGURE 21.9 Iron–cyanide precipitation treatment system process diagram.

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FeSO4 Lime

Acid

SP Leachate with CN, F, metals Reactor 1

Reactor 2

Polymer

Flash Mix/Floc tank

Settling tank

Effluent

Sludge for disposal

FIGURE 21.10 Iron–cyanide precipitation process for treatment of spent potlining leachate at an aluminum production facility.

which the aqueous medium is usually a complex wastewater whose chemical behavior is difficult to predict accurately. An Alcoa plant for treatment of groundwater contaminated with iron cyanide associated with historically disposed spent potlining leachate from aluminum production waste provides a useful example of precipitation process optimization and performance [37]. Spent carbon potlining from aluminum reduction cells at aluminum production facilities was historically landfilled. Spent potlining leachate has a high pH (9 to 12) and elevated concentrations of a number of chemical species, most prominently fluoride (1,200 to 8,500 mg/l), cyanide (150 to 1,250 mg/l), and total dissolved solids (20,000 to 70,000 mg/l) [38]. The cyanide is primarily in the form of dissolved ferrocyanide 3− [Fe(CN)4− 6 ] and ferricyanide [Fe(CN)6 ]. At the Alcoa facility, groundwater contaminated with spent potlining leachate is pumped from the subsurface and the iron–cyanide species are precipitated as Prussian Blue and/or Turnbull’s Blue in an above-ground precipitation treatment process. As shown in Figure 21.10, the pH of the contaminated groundwater is first raised with lime (Ca(OH)2 ) to precipitate fluoride as CaF2 (s), followed by lowering of the pH and addition of ferrous sulfate (FeSO4 ) to precipitate the iron cyanide (see Equation [21.12]). Addition of ferrous iron, Fe2+ , is preferred over Fe3+ in order to minimize precipitation of Fe(OH)3 (s) as a competing reaction, and to promote formation of the low solubility ferric ferrocyanide solids, Prussian Blue and Turnbull’s Blue. There is no two-stage separation; the CaF2 (s) and the iron–cyanide solids are settled simultaneously. Once formed, the CaF2 (s) particles dissolve sufficiently slowly that redissolution of the fluoride upon addition of the sulfuric acid is not substantial within the time period allowed for settling, and in any case CaF2 (s) still has relatively low solubility at mid-range pH. Experiments were conducted to investigate the influence of several process variables on the performance of the iron–cyanide precipitation process. The factors examined were Fe:CN dosing ratio, final pH, and oxidation state of dissolved iron cyanide in the leachate. In the iron–cyanide precipitation process, ferrous sulfate is usually employed as the source of excess iron [24,37,39]. Based on the reaction in Equation (21.12), in which excess ferrous iron is added to react with ferricyanide to precipitate Prussian Blue, the stoichiometric requirement of iron to cyanide on a molar basis is 3:2, which corresponds to a weight basis of 1.3:1. Of course, the ferrous iron and the ferricyanide participate in other reactions, including the redox conversions of the ferrous iron and the ferricyanide, all of which may proceed at different rates under different solution conditions. For these reasons, optimal Fe:CN dosing ratio for precipitating iron cyanide in spent potlining leachate (total cyanide = 33.2 mg/l) was investigated empirically. Bench-scale studies were conducted in which lime was first added to raise the pH and precipitate CaF2 (s), followed by 30 min of rapid stirring. Next, sulfuric acid was added to lower the pH to 7.0 or 6.5, followed by addition of ferrous sulfate and then 30 min of rapid stirring followed by 60 min of slow stirring. Stirring was then ceased

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1.0 pH = 6.5

Total CN (mg/l)

0.8

Reagent sequence = Lime-acid-FeSO4 Total CN, initial = 33 mg/l

pH = 7.0

0.6

1.0

0.4

0.2

0.0 9:1

20:1

40:1

Fe:CN ratio

FIGURE 21.11 Total cyanide remaining in solution vs. ferrous iron:cyanide dosing ratio (weight basis) for lime–H2 SO4 –FeSO4 treatment of spent potlining leachate (total cyanide = 33.2 mg/l) with final pH = 6.5 and 7.0. (Source: Dzombak, D.A. et al., Proceedings of the WEFTEC 96, Water Environment Federation, p. 107, 1996.)

and the sample was allowed to settle quiescently for 18 to 24 h. At the end of this period, supernatant samples were taken for analysis of total cyanide. Results for three different iron:cyanide dosing ratios are shown in Figure 21.11, where it may be seen that dosing of ferrous iron at Fe:CN ratios as low as 9:1 (weight basis) were effective for removal of total cyanide to concentrations below 1.0 mg/l. Additional experiments were performed with the same spent potlining leachate to investigate the effect on total cyanide removal of lowering the final pH to 6.5 rather than 7.0. Formation of Prussian Blue is more thermodynamically favorable at lower pH values [31], and pH 6.5 is near typical lower pH limits specified in discharge permits. Results for tests at pH 6.5 and different iron:cyanide dosing ratios are given in Figure 21.11. As shown there, the total cyanide removals achieved at pH 6.5 were essentially the same as those obtained at pH 7.0. The residual total cyanide concentrations were again lower than 1.0 mg/l for iron:cyanide dosing ratios ranging from 9:1 to 40:1 on a weight basis. The effect of oxidation state of the dissolved iron–cyanide in the untreated spent potlining leachate 4− was investigated by spiking samples with Fe(CN)3− 6 and Fe(CN)6 in different proportions. The total spike amount was 100 mg/l as cyanide in each case. The ferrous iron:cyanide dosing ratio was 9:1 (weight basis) and the final process pH was 7.0 in all cases. Results are presented in Table 21.3, where it may be seen that removal of cyanide was essentially the same in solutions with 100% 4− Fe(CN)3− 6 spike and those with 100% Fe(CN)6 spike. The similar total cyanide removals obtained indicate little effect of the oxidation state of the iron in the leachate iron–cyanide, perhaps because of the large doses of excess ferrous iron added in the treatment process, compared to the relatively small amounts of dissolved iron–cyanide in the leachate. X-ray diffraction analyses performed on precipitate samples from the bench scale tests with the spent potlining leachate all showed the presence of Prussian Blue, or iron–cyanide solids such as Turnbull’s Blue, and Berlin Green, which give the same x-ray diffraction pattern as Prussian Blue [31]. These results provided confirmation that removal of cyanide in the tests was by precipitation, and not by adsorption onto the CaF2 (s) and Fe(OH)3 (s) particles observed to be present. The x-ray diffraction results do not provide sufficient information, however, to identify the specific kind(s) of iron–cyanide solids formed. The bench-scale test results indicate that addition of ferrous sulfate to solutions containing relatively high concentrations (>100 mg/l) of dissolved iron–cyanide species can yield total cyanide

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TABLE 21.3 Effect of Initial Fe(II)–CN/Fe(III)–CN Ratio in Spent Potlining Leachate on Oxidation–Reduction Potential and Total Cyanide Removal by Precipitation of Iron-Cyanide Solid with Ferrous Irona,b Fe(II):Fe(III) in CN spike 0:1 0.5:0.5 1:0

Initial ORP (mV)

Final ORP (mV)

Initial CNT (mg/l)c

Final CNT (mg/l)

+325 +237 +190

−205 −122 −184

133 133 133

0.33 0.30 0.35

a Source: Data from Dzombak, D.A., Dobbs, C.L., Culleiton, C.J., Smith, J.R., and

Krause, D., Proceedings of WEFTE96, Vol. 3, Part I. Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1996, p. 107. b Fe:CN dose ratio = 9:1 (w:w), pH = 7.0. c 100 mg/l CN spike (solutions of Fe(CN)3− and Fe(CN)4− ) added to leachate T 6 6 containing 33 mg/l CNT .

concentrations below 1 mg/l. The consistent residual cyanide concentrations observed reflect the solubility limit of the iron–cyanide solids formed in the pH range studied (equilibrium solubility in the range of 0.5 to 2 mg/l according to Ghosh et al. [31]), and provide further evidence for a precipitative removal mechanism. Full-scale plants have been designed and constructed for removal of iron–cyanide species from contaminated water by precipitation at a number of aluminum smelting facilities. At the Alcoa facility mentioned above, for example, an average flow of 36,000 gallons/day (25 gpm) of groundwater contaminated with spent potlining leachate is treated continuously. The treatment scheme is as shown in Figure 21.10. Some example process performance data are given in Figure 21.12, where it may be seen that influent total cyanide concentrations ranging from 15 to 45 mg/l were consistently reduced by ferrous sulfate addition and iron–cyanide precipitation to levels of 0.1 to 0.2 mg/l. The iron:cyanide mass dosing ratio employed in the full-scale process during the operation period represented in Figure 21.12 was 36:1. This high dosing ratio was employed by the plant operators to provide a factor of safety. As indicated by the bench-scale data in Figure 21.11, however, the same degree of cyanide removal could likely be achieved with a lower iron:cyanide dosing ratio. Iron–cyanide precipitation has also been investigated for treatment of cyanide in electroplating wastewaters. Hassan et al. [24] treated electroplating wastewater containing 350 mg/l total cyanide, of which most was present as free cyanide (33 mg/l was strongly complexed and not amenable to destruction by chlorination). A two-stage precipitation process was employed, with addition of ferrous sulfate, FeSO4 , at pH 8.5, followed by pH adjustment to pH 6.5 and addition of ferric chloride, FeCl3 . The rationale for the ferrous and ferric iron salt additions at two different pH values was not provided, though presumably, it relates to maintaining solubility of Fe2+ and Fe3+ . Ferric iron is more soluble at pH 6.5 than at pH 8.5, but this is also the case for Fe2+ , as oxidation kinetics are slower and ferrous carbonate and ferrous hydroxide are less likely to form. The value of the two-stage salt addition is thus unclear. Nevertheless, with the two-stage addition of ferrous and ferric iron, a solid described as Prussian Blue (Fe4 (Fe(CN)6 )3 (s)) was formed, resulting in reduction of total cyanide in solution from 350 to 83.0 mg/l. Concentrations of various metals present in the electroplating wastewater were also reduced somewhat, as indicated in Table 21.4. The doses of FeSO4 and FeCl3 that yielded these results were not provided. It is unclear as to whether higher doses or a different dosing regime could have yielded lower residual cyanide in solution. Analyses of the sludge produced, also given in Table 21.4, showed very high concentrations of cyanide, and high concentrations of the metals removed from solution. As indicated earlier in the chapter,

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Reagent sequence = Lime-acid-FeSO4 Fe:CN ratio = 36:1 50 Influent

Total CN (mg/l)

40 30 20

Effluent (0.1 to 0.2 mg/l)

10 0 25-Aug

26-Aug

27-Aug

28-Aug

29-Aug

30-Aug

31-Aug

Date in August, 1996

FIGURE 21.12 Influent and effluent total cyanide concentrations for a full-scale precipitation process (see Figure 21.9) treating spent potlining leachate at an aluminum production facility. Fe:CN ratio = 36:1. pH range 6.6 to 6.7. Flow range = 25 to 36 gpm. Courtesy Alcoa, Inc.

TABLE 21.4 Results of Treatment of Electroplating Wastewater with FeSO4 and FeCl3 to Precipitate Free Cyanide as Iron-Cyanide Solidsa CNT (mg/l)

NATCb CN (mg/l)

Cd (mg/l)

Cu (mg/l)

Ni (mg/l)

Zn (mg/l)

Fe (mg/l)

Raw waste Fe–CN Ppt. eff.

350 83.0

33.0 73.0

29.5 0.70

68.7 55.8

14.8 13.6

12.2 0.15

13.3 1.93

Fe–CN sludge

15,600

15,500

106

567

62,600

Sample

(mg/kg wet sludge) 1690 53.0

a Source: Adapted from Hassan, S.Q., Vitello, M.P., Kupferle, M.J., and Grosse, D.W., J. Air Waste Mange. Assoc., 41, 710, 1991. b NATC = not amenable to chlorination.

mixed-metal–cyanide solids can form, and metals can exchange with Fe2+ and Fe3+ in iron–cyanide solids. In addition, at pH 6.5 and 8.5 the formation of ferric hydroxide, Fe(OH)3 (s) is favored, and thus the solids formed may have been a mixture of Prussian Blue and ferric hydroxide, and perhaps a solid solution which can occur when excess ferric iron is present [31]. Ferric hydroxide has significant adsorptive capacity for metals [30].

21.2.3 OPERATION AND REGULATORY ISSUES Iron–cyanide precipitation is a robust technology for treatment of cyanide-bearing waters in that it can be applied to remove cyanide from very complex waters. It has been employed in treating wastewaters from steel and aluminum manufacturing, electroplating, and other industries. The technology has been shown capable of reducing high concentrations of free or complexed cyanide to concentrations in the range of 1 to 10 mg/l. Process schemes for implementation are simple, involving reagent addition and mixing, and settling of the precipitated solids.

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While iron–cyanide precipitation is a simple, robust treatment technique, it has limitations common to all precipitation processes. The contaminant removals achievable are fundamentally limited by the solubility of the solids formed. In the case of iron–cyanide solids, minimum solubilities are in the range of several mg/l, which may not be sufficient to meet discharge limits in some cases. Additional treatment, for example, by ion exchange, may be necessary. A further limitation of the process is that a sludge is generated that will usually require management as a hazardous waste. This adds significantly to the cost of the process. In some cases, the sludge generated may not be eligible for land disposal, even as a hazardous waste, due to its characteristics or regulatory designation. In such cases the sludge will require thermal or other aggressive treatment to render it non-hazardous, raising the cost of sludge management. An example of the potential complexities of sludge management for sludge from iron–cyanide precipitation processes, is provided by the use of this technique for treatment of spent potlining leachate at aluminum manufacturing facilities. Spent potlining is a “listed” hazardous waste (K088) under RCRA hazardous waste regulations in the United States, and by the “derived-from rule” in the regulations (40 CFR 261.3 (d) (2)), the leachate and any residuals generated in its treatment are also listed hazardous wastes. Hence, iron–cyanide sludge from precipitation treatment of spent potlining treatment is a listed hazardous waste and must be managed as such. In the United States, this means that the waste cannot be land disposed directly if the total cyanide concentration in the solid exceeds the Land Disposal Restriction (LDR) standard of 590 mg/kg (see Chapter 18). In that case, it may require thermal or other treatment and meet specified post-treatment characteristics in order to be landfilled.

21.3 AIR STRIPPING Since hydrogen cyanide, HCN, is a volatile compound, its tendency to partition from water to air can be exploited to remove it from water. In this approach, water-containing cyanide is introduced to a reactor in which air flow is also introduced and moves through the water, resulting in transfer of HCN from water to air. As HCN has moderate volatility, and the energy cost of blowing air through an air stripper is significant, air stripping is generally not used to treat cyanide-bearing waters to meet discharge requirements. Air stripping can be useful, however, for reducing free cyanide concentrations in solutions and slurries to levels readily treatable by more cost effective technologies such as biological oxidation.

21.3.1 COUNTER-CURRENT AIR STRIPPING Air stripping of contaminants from water is most often performed using counter-current, packed tower air stripping [40,41]. In this process, water containing the volatile contaminant is introduced at the top of a column containing high-surface-area, porous plastic, or ceramic packing material. Air is introduced at the bottom of the column, and flows upward through the column as the water trickles down through the media to a collection sump at the bottom of the tower. As the counter-current flows of air and water occur, the volatile species is transferred from the water to the air stream, and exits with the air flow at the top of the column. The treated water exits at the bottom of the tower. Because of the moderate volatility of HCN and its toxicity, air stripping is not the leading treatment approach for cyanide-contaminated water in most cases. The moderate volatility of HCN at ambient temperatures means that high air:water ratios are required, resulting in relatively high capital and operating costs [41]. The toxicity of HCN means that direct atmospheric discharge of air stripper off-gas from systems containing appreciable amounts of cyanide is not feasible. For waters containing cyanide, off-gas from air strippers must be treated. Hydrogen cyanide is removed in the ammonia steam stripper in coke plant wastewater treatment (see Chapter 26), and the stripper offgas is generally returned to the coke oven gas (COG) main pipes. COG is often treated for recovery

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of ammonia, sometimes desulfurized, and then combusted. In desulfurization, cyanide is generally removed from the COG as thiocyanate. For the more common ambient-temperature air stripping, offgas is usually treated by passing the gas flow through granular activated carbon [42]. HCN adsorbs on activated carbon, as discussed earlier in the chapter, and can be recovered. In the gold mining industry, the stripped HCN is absorbed into a solution of NaOH to about 5% cyanide by weight; this solution is returned to the heap leach process [43,44].

21.3.2 RECOVERY AND REUSE OF CYANIDE IN HYDROMETALLURGICAL GOLD MINING Cyanide recovery by air stripping has been done in the mining industry since about 1930. Mudder and Goldstone [44] present a concise history of the use of air stripping for cyanide recovery. In the early days, recovery was economically driven because of the limited production of sodium cyanide. Today, recovery is driven by both economics and environmental concerns. The cost of using cyanide in gold recovery is at least $3.00/ton of ore (cyanide usage of 2.5 kg/ton of ore processed and cyanide cost of $1.20/kg). It has been estimated that 20 to 30% of these costs can be saved by implementing cyanide recovery [43]. Cyanide can be recovered from several processing streams in hydrometallurgical gold mining, including barren solution and tailings, tailings slurry, pregnant solution, and mill slurry before cyanidation. Recovery of cyanide from pregnant liquor prior to CIP (carbon in pulp) processing has been shown to increase gold recovery by reducing the competition of compounds, other than the gold complex, for active sites in carbon adsorption (see Section 21.1.1); cyanide is also a competing compound for adsorption sites [43]. Another reason for recovering cyanide prior to CIP is that cyanide oxidation by oxygen from the atmosphere is catalyzed by activated carbon and can be a significant source of cyanide losses [43]. The most common cyanide recovery process is an air or vapor stripping process. It includes the following processing stages as shown in Figure 21.13. • Acidification to pH 5 to 8 to convert ionic or weakly complexed cyanide to non-dissociated HCN. • Hydrogen cyanide stripping followed by adsorption into a sodium hydroxide solution. • Reneutralization of stripped tailings with lime to precipitate metals. • Clarification (if required). This step is not required for barren slurries, which can be charged directly to a tailing pond. Several processing factors impact the effectiveness of cyanide air stripping. First, the solution or slurry characteristics can have a significant impact on the recovery process. The concentration of cyanide and the quantity of solution or slurry dictate the size of the processing facility. Cyanide speciation may make recovery infeasible since the predominance of strong metal–cyanide complexes inhibits recovery. On the other hand, predominance of WAD cyanide favors recovery. Solution pH, which can be adjusted, temperature, cyanide concentrations, and cyanide speciation combine to dictate the production of HCN, the strippable cyanide compound. In addition, solution or slurry viscosity affects process equipment selection, and the efficiency of the mass transfer of HCN from solution phase to the gas phase. Two process configurations, complete-mix reactor tank (Figure 21.14) and packed-bed tower (Figure 21.15), have been considered for recovery of HCN from process streams. Packed towers offer more efficient cyanide stripping capabilities by providing multiple mass transfer stages for the transfer of the cyanide from the aqueous to the gas phase. In contrast, a complete-mix reactor unit only provides one mass transfer stage; if several stages are required to achieve desired performance, several of these reactors will be required. A packed tower also offers the benefit of smaller space requirement. While there can be operating problems with gypsum or calcite scale-formation in packed towers, these scaling issues can be managed using anti-scaling agents.

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Barren water or tailings slurry

Air (mixing)

Concentrated sulfuric acid

Acidification (pH 5 to 8)

NaOH

Air

Cyanide stripping

(HCN)

Absorption

NaCN to process Reneutralization for metals precipitation

Flocculant (if required)

Clarification

Lime addition to pH = 9.5

Slurry to tailings pond

Tertiary treatment (if required for discharge)

FIGURE 21.13 Schematic diagram of air stripping cyanide recovery process in gold mining. (Source: Mudder, T. and Goldstone, A., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

The primary process operating parameters for a packed tower are air-to-water ratio (v/v), air/water contact time, and temperature. Unless inexpensive waste heat is available, the process operating temperature will be that of the solution or slurry that is delivered for processing, which is usually at or near ambient air temperature. The air-to-water ratio and air/water contact time are the two main process operating parameters and they determine the size of the stripping tower and the effectiveness of the HCN stripping. Because of the moderate cyanide concentrations in the gold mining process streams (hundreds of milligrams per liter as opposed to thousands) and the moderate volatility of HCN, air requirements to the reactor tend to be high. For low aqueous phase concentrations, the equilibrium gas concentration will be commensurately low. The achievement of the target concentration will dictate the necessary air/water contact time. For equivalent stripping efficiency, a complete-mix reactor will require longer contact time and greater air-to-water ratios than will a packed tower system. Packed media towers operate on the basis of mass transfer through a liquid film on the packing media as opposed to bulk liquid media transfer at an air bubble interface, which is what takes place in a complete-mix reactor (see Figure 21.16). Liquid film mass transfer requires a shorter contact time than does bulk liquid mass transfer. Economically efficient cyanide recovery by air stripping will not produce treated water quality that complies with discharge requirements. Recovery limits are established to balance the costs for

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HCN air H2O

H2SO4

pH Solution or slurry

AIR

AIR

AIR HCN strippers

Acidification section Air

Air (to atmos.)

Lime

NaOH

pH

HCN scrubbing recovery section Tailings thickener/ disposal

Reneutralization section

Mill circuit

5% NaCN solution

FIGURE 21.14 Schematic diagram of a complete-mix-air-stripping cyanide recovery system. (Source: Adapted from Mudder, T. and Goldstone, A., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

recovery against the cost for final polishing treatment, for example, added cost for greater air-towater ratio to increase recovery versus incremental cost to treat additional cyanide using an oxidant like hydrogen peroxide. The primary commercial technologies used in gold mining cyanide recovery are as follows: • Mills Crow process — Original process, used at the Flin Flon Mine in Canada from about 1930 to 1975; involved acidification, air stripping in packed towers, and reabsorption in lime slurry sprays [44–46]. • AVR (acidification, volatalization, and reneutralization) process — CANMET. A modification of the Mills Crow process, utilizing single complete-mix aeration reactors for cyanide stripping [43]. • Cyanisorb process — Utilizes packed medium stripping and adsorption towers. Figure 21.17 presents a process flow diagram for the full-scale Cyanisorb slurry tailings processing plant at the Golden Cross Mine near Waihi, New Zealand [43]. • Air-Sparged Hydrocyclone — Counter-current flow through vertical tubes with porous inner tube; solution or slurry is fed through cyclone header to create high-velocity swirl flow, and air is introduced along the tubes with fine bubbles created by the induced fluid shear [46]. Table 21.5 presents a comparison of performance of these processes in recovering cyanide. Performance will vary with changes in operating conditions, but the data in Table 21.5 serve to illustrate that high recoveries of cyanide can be achieved by air stripping.

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HCN air H2O H2SO4

Solution or slurry AIR

HCN strippers

AIR

Acidification section Air

Air (to atmos.)

Lime

NaOH

HCN scrubbing recovery section Tailings thickener/ disposal

Mill circuit

5% NaCN solution

FIGURE 21.15 Schematic diagram of a packed tower air stripping cyanide recovery system. (Source: Adapted from Mudder, T. and Goldstone, A., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

Water film CW

Bulk air flow CW,i

Conc.

CA-bulk air concentration Distance

Packing medium surface

FIGURE 21.16 Illustration of mass transfer from a liquid to a gas phase, and effect of liquid film.

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1.2 to 1.5 kg H2SO4/M3 tailings slurry

213 M3/hr 224 mg/l WAD pH 10 to 11

Air/liquid = 330 to 1000

Air

Air

Acidification pH 7 to 8 tank 10 to 15 min

Tailings slurry Tailings slurry

Lime

Stripper bottoms tank Tallings pond

Return to CIP Neutralization

pH 9.0 to 9.5

31 mg/l WAD

Recovered cyanide

50.5 kg Na CN/h

FIGURE 21.17 Cyanisorb cyanide recovery process flow diagram, Gold Cross Mine, Waihi, New Zealand. (Source: Goldstone, A. and Mudder, T., Cyanide Monograph, Mining Journal Books, Ltd., London, 1998. With permission.)

TABLE 21.5 Performance Comparison of Cyanide Recovery Processes used in Hydrometallurgical Gold Mininga Recovery technology Mills-Crowe AVR (Canmet) Cyanisorb ASH ASH

Reactors employed 4 towers 2 towers 2 towers 1 ASH 1 ASH

Streams treated

Air:water ratio

CNin (mg/l)

CNout (mg/l)

% CN recovered

521 330 NAb 100 10

560 330 600 250 250

44 2 30 25 50

92 99 95 90 80

solution solution slurry slurry solution

a Source: Data from Parga, J.R. and Miller, J.D., Cyanide: Industrial and Economic Aspects, Young, C.A., Twidwell, L.G.,

and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrandale, PA, 2001. b NA: no data available.

21.4 SUMMARY AND CONCLUSIONS • Physical–chemical separation of cyanide species from water by adsorption and precipitation is commonly used to treat waters with metal–cyanide species (e.g., leachate

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• • •

• • • •

•

435

from historically disposed spent potliner material generated in aluminum manufacturing, groundwater at former manufactured gas plant sites, and iron and steel process wastewaters). Various adsorbents have been demonstrated at bench, pilot, and full-scale to be useful for removal of cyanide species from water by fixed-bed treatment, including granular activated carbon, synthetic anion exchange resins, and activated alumina. Granular activated carbon is the most widely used adsorbent for fixed-bed treatment of cyanide in water. It has been shown to be effective in removing both free cyanide and metal–cyanide species. Anion exchange has been demonstrated to be capable of removing cyanides, especially metal–cyanide complexes, from aqueous solution, but has seen limited use because of the relatively high cost of ion exchange materials, limitations in regeneration efficiency, and concern about HCN gas formation during resin regeneration. Granular activated alumina is effective for removing metal cyanides, but has seen little use for treatment of cyanide-bearing waters because, activated carbon and ion exchange technologies are more commercially accessible. Precipitation has been employed primarily for removal of metal cyanides, especially strong metal–cyanide complexes like iron cyanides, and for waters with high cyanide concentrations, that is, greater than 50 mg/l. Iron-cyanide precipitation, the most commonly employed precipitation process, is a robust technology for treatment of cyanide-bearing waters in that it can and has been applied to remove cyanide from very complex waters. Iron-cyanide precipitation has limitations common to all precipitation processes: contaminant removals achievable are limited by the solubility of the solids formed, with minimum solubilities for iron–cyanide solids in the range of several mg/l; and the sludge generated will require expensive management, usually as a hazardous waste in the case of iron–cyanide sludges. Air stripping can be used to recover free cyanide from aqueous solutions and slurries. Full-scale air stripping units have been in operation for many years at hydrometallurgical gold mining operations for cyanide recovery. Due to the moderate volatility of HCN, required air-to-water ratios are high for removal of cyanide, which has limited more widespread application of air stripping for HCN removal from water.

REFERENCES 1. Theis, T.L. and West, M.L., Effects of cyanide complexation on the adsorption of trace metals at the surface of goethite, Environ. Technol. Lett., 7, 309, 1986. 2. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 3. Guo, R., Chakrabarti, C.L., Subramanian, K.S., Ma, X., Lu, Y., Cheng, J., and Pickering, W.F., Sorption of low levels of cyanide by granular activated carbon, Water Environ. Res., 65, 640, 1993. 4. Chatwin, T.D., Zhang, J., and Gridley, G.M. Natural mechanisms in soil to mitigate cyanide release, in Proceedings Superfund ’88, The 9th National Conference, Hazardous Materials Control Research Institute, Washington, DC, 1988, p. 467. 5. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sediment Contam., submitted, 2005. 6. Jay, W.H., Copper–gold cyanide recovery systems, in Cyanide: Social, Industrial and Economic Aspects, Young, T.C, Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 317. 7. Young, C.A., Remediation technologies for the management of aqueous cyanide species, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 175.

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8. Patterson, J.W., Cyanide, in Industrial Wastewater Treatment Technology, 2nd ed., Butterworth, Boston, MA, 1985, p. 115. 9. Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Surprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Noyes Data Corp., Park Ridge, NJ, 1988. 10. Reed, A.K., Shea, J.F., Tewksbury, T.L., Cherry, R.H., and Smithson, G.R., An investigation of techniques for removal of cyanide from electroplating wastes, 12010 EIE, U.S. Environmental Protection Agency, Office of Research and Monitoring, Washington, DC, 1971. 11. Huff, J.E., Fochtman, E.G., and Bigger, J.M., Cyanide removal from refinery wastewater using powdered activated carbon, in Carbon Adsorption Handbook, Cheremisinoff, P.E. and Ellerbusch, F., Eds., Ann Arbor Science, Ann Arbor, MI, 1978, p. 733. 12. Bernardin, F.E., Cyanide detoxification using adsorption and catalytic oxidation on granular activated carbon, J. Water Poll. Control Fed., 45, 221, 1973. 13. Aksu, Z. and Calik, A., Adsorption of iron(III)-cyanide complex ions to granular activated carbon, J. Environ. Sci. Health, A34, 2087, 1999. 14. Saito, I., The removal of hexacyanoferrate(II) and (III) ions in dilute aqueous solution by activated carbon, Water Res., 18, 319, 1984. 15. Chen, Y.S., You, C.G., and Ying, W.C., Cyanide destruction by catalytic oxidation, Metal Finish., 89, 68, 1991. 16. Kuhn, R.G., Process for detoxification of cyanide-containing aqueous solutions, U.S. Patent 3,586,623, June 22, 1971. 17. Chank, J.K., pH-dependent adsorption of hexacyanoferrate(II) onto selected sorbents, M.S. thesis, Clarkson University, Potsdam, NY, 1997. 18. Mintek, The carbon-in-pulp process, www.mintek.ac.za/EMD/CIP/cip.htm, accessed: August 28, 2003, 19. Botz, M. and Mudder, T., Mine water treatment with activated carbon, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 20. Gangloff, C., Environmental concerns driving activated carbon market, Water Technol., May, http://www.watertechonline.com/article.asp?IndexID=5220504, 1999. 21. Tran, T., Fernando, K., Lee, K., and Lucien, F., Use of ion exchange resin for the treatment of cyanide and thiocyanate during the processing of gold ores, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds. The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 289. 22. Leao, V.A., Ciminelli, V.S.T., and Costa, R.D.S., Cyanide recycling using strong-base ion exchange resins, JOM, 50, 71, 1998. 23. Lukey, G.C., Van Deventer, J.S.J., and Shallcross, D.C., The effect of functional group structure on the elution of metal cyanide complexes from ion-exchange resins, Separation Sci. Technol., 35, 2393, 2000. 24. Hassan, S.Q., Vitello, M.P., Kupferle, M.J., and Grosse, D.W., Treatment technology evaluation for aqueous metal and cyanide bearing hazardous wastes (F007), J. Air Waste Manage. Assoc., 41, 710, 1991. 25. Clifford, D.A., Ion exchange and inorganic adsorption, in Water Quality and Treatment, 5th ed, Letterman, R.D., Ed., American Water Works Assoc., McGraw-Hill, New York, NY, 1999, p. 9.1. 26. Vachon, D.T., Removal of iron cyanide from gold mill effluents by ion exchange, Water Sci. Technol., 17, 313, 1985. 27. Moore, F.L., Improved ion-exchange resin method for removal and recovery of zinc cyanide and cyanide from electroplating wastes, J. Environ. Sci. Health, Part A, 11, 459, 1976. 28. Mullin, R., Basic materials keep a technology edge, Chem. Eng. News, 80, 44, 2002. 29. Bushey, J.T. and Dzombak, D.A., Ferrocyanide adsorption on aluminum oxides, J. Coll. Int. Sci., 272, 46, 2004. 30. Dzombak, D.A. and Morel, F.M.M., Surface Complexation Modeling: Hydrous Ferric Oxide, WileyInterscience, New York, NY, 1990. 31. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 32. Haas, P.A., A review of information on ferrocyanide solids for removal of cesium from solutions, Separation Sci. Technol., 28, 2479, 1993. 33. Tanihara, K., Yasuda, S., and Nishikubo, K., Precipitation of iron cyanide complexes from cyanide zinc electroplating, Metal Finish., 85, 131, 1987.

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34. Adams, M.D., Precipitation of cyanide as Cu2 Fe(CN)6 compounds from cyanidation and detoxification circuits, in Proceedings of the International Congress on Mineral Processing and Extractive Metallurgy, Australian Institute of Mining and Metallurgy, Melbourne, Australia, 2000, p. 201. 35. Sierka, R.A., Complexed cyanide and fluoride treatment by chemical coagulation and membrane processing, 95-02-02, Environmental Technology Center, Alcoa Technical Center, Alcoa, Inc., Alcoa Center, PA, 1995. 36. Sinha, P.K., Amalraj, R.V., and Krishnasamy, V., Flocculation studies on freshly precipitated copper ferrocyanide for the removal of cesium from radioactive liquid waste, Waste Manage., 13, 341, 1993. 37. Dzombak, D.A., Dobbs, C.L., Culleiton, C.J., Smith, J.R., and Krause, D., Removal of cyanide from spent potlining leachate by iron cyanide precipitation, in Proceedings of WEFTEC96, Vol. 3, Part I. Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1996, p. 107. 38. Blayden, L.C., Hohman, S.C., and Robuck, S.J., Spent potliner leaching and leachate treatment, in Proceedings of Light Metals 1987, The Minerals, Metals and Materials Society, Warrendale, PA, 1987, p. 663. 39. ORSANCO, Methods for treating metal-finishing wastes, Ohio River Valley Water Sanitation Commission, Cincinnati, OH, 1953. 40. Nyer, E.K., Groundwater Treatment Technology, 2nd ed., Von Nostrand Reinhold, New York, 1992. 41. Dzombak, D.A., Roy, S.B., and Fang, H.J., The air stripper design and costing (ASDC) program, J. Am. Water Works Assoc., 85, 63, 1993. 42. Adams, J.Q. and Clark, R.M., Evaluating the costs of packed-tower aeration and GAC for controlling selected organics, J. Am. Water Works Assoc., 83, 49, 1991. 43. Stevenson, J., Botz, M., Mudder, T., Wilder, A., Richins, R., and Burdett, B., Recovery of cyanide from mill tailings, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 44. Mudder, T. and Goldstone, A., Recovery of cyanide from slurries, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 45. Botz, M. and Stevenson, J., Cyanide: recovery and destruction, Eng. Mining J., June, 44, 1995. 46. Parga, J.R. and Miller, J.D., Cyanide recovery/destruction using air sparged hydrocyclone technology, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 363.

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and High Temperature 22 Thermal Oxidation Technologies for Treatment of Cyanide

Rajat S. Ghosh, John R. Smith, and George M. Wong-Chong CONTENTS 22.1

22.2

22.3

22.4

22.5

22.6

High Temperature Alkaline Hydrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.1.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.1.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.1.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.1.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.1.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . High Temperature Alkaline Chlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.2.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Incineration/Thermal Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.3.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.3.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.3.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.3.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.3.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Electrolytic Decomposition or Oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.4.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.4.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.4.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.4.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.4.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Polysulfide Process. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.5.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wet Air Oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.6.1 Process Description and Implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.6.2 Achievable Treatment Levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.6.3 Technology Selection Factors and Design Considerations. . . . . . . . . . . . . . . . . . . . . . 22.6.4 Cost of the Technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

440 440 442 442 443 443 443 443 444 444 444 445 445 445 446 446 446 446 447 447 447 447 448 448 448 448 449 449 450 450 450 450 451 451 452 439

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22.6.5 Technology Status. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Other Thermal Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.7.1 Thermal Desorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.7.2 Thermal Oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.7.3 Thermal Plasma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.8 Thermal Technology Overview. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.9 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.7

452 452 452 453 453 453 456 456

Certain thermal technologies can be used to destroy all forms of cyanide, including the strong metal–cyanide complexes (e.g., iron and cobalt) in both solid and liquid wastes. Generally these technologies involve both high temperature and pressure, and they tend to be relatively expensive to implement and operate when compared to ambient-temperature treatment options (e.g., adsorption, ion exchange, and chemical precipitation processes — see Chapters 20, 21, and 23). However, given the right set of circumstances, these thermal technologies can be cost effective, for example, when dealing with small quantities of waste and off-site thermal treatment facilities are available. The following thermal technologies are described in detail in the following sections: • • • • • •

High temperature alkaline hydrolysis High temperature alkaline chlorination Incineration or thermal treatment Electrolytic oxidation or (decomposition) Calcium polysulfide treatment Wet air oxidation

These technologies have been applied for the treatment of water, soil slurries, and sludges containing free cyanide, weak metal–cyanide complexes, and strong metal–cyanide complexes. Descriptions for the technologies follow, and include the following main features: • Process description and implementation • Achievable treatment levels • Design considerations • Critical design conditions • Residuals generated • Technology complexity • Cost information • Status of technology implementation The chapter concludes with a technology summary matrix (Table 22.3) for the thermal technologies discussed here. Only some of the technologies have been demonstrated at full-scale.

22.1 HIGH TEMPERATURE ALKALINE HYDROLYSIS 22.1.1 PROCESS DESCRIPTION AND IMPLEMENTATION High temperature alkaline hydrolysis (HTAH) entails hydrolytic cracking of metal–cyanide complexes at high temperature and pressure under alkaline conditions. The overall breakdown reaction for iron cyanide is as follows: − − 6Fe(CN)4− 6 + 12OH + 66H2 O + O2 → 36NH3 + 2Fe3 O4 (s) + 36HCOO

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This reaction proceeds in a sequence of steps. First, the iron–cyanide complex is broken down liberating free cyanide in an alkaline bath under high temperature and pressure (e.g., in the range of 165–180◦ C and 100–150 psig). The liberated free cyanide is hydrolyzed to formate and ammonia, and at the elevated temperatures and pressures, formate and ammonia can be further oxidized to H2 O, CO2 , and N2 . Robuck and Luthy [1] performed bench-scale testing with spent potlining leachate containing approximately 325 mg/l total iron and 925 mg/l total cyanide (predominantly iron–cyanide complexes). In these studies, total cyanide was reduced to less than 0.05 mg/l and the iron was oxidized to Fe3 O4 (s) particles at the following process conditions: • • • •

Temperature: 177 to 275◦ C Pressure: 10 to 62 atm pH: ∼12.5 Reaction time: 2 h

Table 22.1 summarizes results of treating the spent potlining leachate under various temperature and pressure conditions. The destruction of ferrocyanide was observed to be a first order reaction with a rate constant of approximately 7 × 10−4 sec−1 . The leachate also contained 310 mg/l of thiocyanate (SCN− ) that was not appreciably reduced except at the 275◦ C and 900 psig condition where about 70% reduction was realized. Subsequent to the bench-scale treatability work by Robuck and Luthy [1], a pilot-scale unit was set up by Kimmerle et al. [2] to treat free cyanide and ferrocyanides (free cyanide ≤20 ppm; ferrocyanide ∼4000 ppm) in spent potlining leachate and a slurry of spent potliner solids. The results from this study indicated 99.9% destruction of cyanide after only 30 min at a temperature of 200◦ C. Osaka Gas Company built a full-scale enhanced alkaline hydrolysis system to treat iron–cyanideimpacted wastewaters from a nitriding process [3]. A process flow diagram is provided in Figure 22.1. The raw wastewater containing approximately 14,000 mg/l of total cyanide (predominantly iron– cyanide complexes) and 2,300 mg/l total iron was treated down to a soluble iron level of less than 4 mg/l and a total cyanide concentration below 0.1 mg/l using a two-step reaction process. In the first

TABLE 22.1 High-Temperature Alkaline Hydrolysis of Cyanide in Spent Potlining Leachate at Various Temperatures and Pressures

Parameter pH CN-total Formate NH3 SCN− Fe-total N-total

Operating temperature, ◦ C (pressure, psig)

Initial conditions (mg/l)

275 (900)

252 (630)

232 (440)

215 (310)

192 (200)

177 (135)

143 (49)

12.7 units 925 212 1230 310 305 1720

12.5 <0.05 1700 1940 88 <0.1 1620

—a <0.05 1730 1910 — 0.1 1620

12.6 <0.05 1690 1900 265 <0.1 1640

— <0.05 1740 1800 274 — —

— <0.05 1760 1850 279 — —

12.6 <0.05 1750 1870 292 — —

— 100 1550 1800 303 — —

a — indicates no data reported.

Source: Reprinted from Water Science and Technology, volume 21, issue number 6–7, Pages 547– 558, with permission from the copyright holders, IWA.

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Exhaust

Alkaline wastewater

Pump Reactor Heat exchanger

Gas–liquid separator Cooler

Compressor Treated water Air

FIGURE 22.1 High temperature alkaline hydrolysis process. (Source: Futakawa, M., et al., Desalinization, 98, 345, 1994. With permission.)

step, iron–cyanide complexes in the raw wastewater were decomposed to formic acid and ammonia in the pretreatment reactor, operated at a temperature of 220◦ C, a pressure of 29 atm, and a residence time of 1 h. In the second step, formic acid and ammonia were further oxidized to water, CO2 , and nitrogen in the catalytic reactor operated at 230◦ C and 39 atm for 1.5 h. Heritage Environmental Services has operated a high temperature/high pressure cyanide destruction unit (CDU) since 1989 [4]. This CDU is operated in batch mode. Cyanide-bearing wastes from various industries (e.g., electroplating, plating, and aluminum smelting) are processed at temperatures in the range of 205–260◦ C (400–500◦ F), pressures in the range of 600–700 psig, and pH in the range of 11–12. The 1,000-gallon batch reactor is charged with waste containing 20–100,000 mg/l total cyanide (on average about 10,000 mg/l), essentially all strongly complexed cyanide, such as iron cyanides. After processing through an 8-h cycle the cyanides are reduced to about 4 mg/l total cyanide. Each operating cycle comprises a 2-h heat-up period, a 4-h reaction period, and a 2-h cool-down period.

22.1.2 ACHIEVABLE TREATMENT LEVELS Bench-scale HTAH processing has shown that total cyanide concentrations can be reduced to less than 0.1 mg/l irrespective of the influent concentrations. This level of treatment greatly depends on reaction conditions (i.e., temperature, pressure, pH, and reaction time) and the presence/absence of other reaction inhibiting constituents, for example, copper.

22.1.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The controlling design parameters for high-temperature alkaline hydrolysis (HTAH) include temperature, pressure, reaction pH, and reaction time. The rate of the alkaline hydrolysis reaction is highly dependent on temperature and pressure as shown by Robuck and Luthy [1]. Since HTAH is largely carried out in commercial pressure vessels, the technology is most well-suited for waste flow rates less than 25 to 30 gpm. Waste streams with pH > 10 and TDS < 1000 ppm are most appropriate for high-temperature alkaline hydrolysis. According to Robuck and Luthy [1], alkaline hydrolysis is a technically viable process for treating small volumes of waste with high concentrations of strong metal–cyanide complexes (>50 ppm of strong metal–cyanide complexes). However, the technology is particularly not suitable for treating waste streams containing copper– cyanide complexes, since this group of complexes exhibits slower decomposition kinetics (an order of magnitude slower than ferrocyanide complex) under the same process conditions.

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The metal in the cyanide complex and that present in the waste will form a hydroxide–oxide sludge that will require management and disposal. Soluble iron in the waste is oxidized to particulate Fe3 O4 (s), which leads to the production of a more compact sludge cake compared to that from a hydroxide sludge. Implementation of the technology requires high-temperature pressure vessels capable of withstanding pressures up to 40 atm. These are available from commercial vendors. Like most other energy-intensive technologies, HTAH needs monitoring to maintain optimum pressure and temperature during the reactor operation as well as routine equipment maintenance and attention to process operating procedures to maintain desired operating conditions. Because this technology is commonly operated in batch mode, low volume, concentrated cyanide wastes are most suitable to treat.

22.1.4 COST OF THE TECHNOLOGY The high-temperature alkaline hydrolysis process can be expensive to implement and operate. The estimated capital costs for a full-scale 25 gpm continuous treatment system for treating spent potlining leachate with high concentration of iron–cyanide complexes (>200 ppm), high pH (10 to 12), high TDS (>20,000 ppm), high carbonates (>6, 000 ppm), high fluoride (∼1,000 ppm), and high Na (∼7,000 ppm) has been estimated (preliminary) as approximately $2.2 million with operation and maintenance costs around $0.44 million/year (2001 cost basis, unpublished data, Alcoa Inc.).

22.1.5 TECHNOLOGY STATUS HTAH has been evaluated in bench-scale studies [1], tested at pilot-scale at an Alcan facility for spent potlining leachate treatment of iron cyanide complexes [2], and implemented in full scale at a chemical plant in Japan and in an aluminum smelting facility in Québec, Canada. However, pilot-scale testing or at least some bench-scale testing is necessary on a case-by-case basis to develop sitespecific design criteria and operating conditions for full-scale treatment of a particular waste stream at a particular site.

22.2 HIGH TEMPERATURE ALKALINE CHLORINATION 22.2.1 PROCESS DESCRIPTION AND IMPLEMENTATION High-temperature alkaline chlorination (HTAC) technology involves the same underlying principles used in the alkaline chlorination process (discussed in Chapter 20), except that it operates at a temperature of approximately 140 to 180◦ C [5,6]. The process involves three steps: (1) the dissociation of the metal cyanide to liberate free cyanide, (2) oxidation of free cyanide to CNO− , and (3) oxidation of CNO− to CO2 , H2 O, and N2 . The high reaction temperature under alkaline chlorination conditions allows dissociation of strong and weak metal–cyanide complexes by the following reaction: − − + − 2Fe(CN)4− 6 + 3OH + 3H2 O → 12CN + 2Fe(OH)3 (s) + 3H + 2e

(22.2)

Once free cyanide is generated, chlorination under alkaline conditions oxidizes free cyanide to cyanate and then cyanate to carbonate and nitrogen by the following set of reactions: Cl2 + CN− + 2NaOH → CNO− + 2Na+ + 2Cl− + H2 O

(22.3)

3Cl2 + 2CNO− + 4NaOH → 2CO2 + N2 + 6Cl− + 4Na+ + 2H2 O

(22.4)

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Similar to the ambient-temperature alkaline chlorination process discussed in Chapter 20, addition of alkali to a pH of approximately 10 during the first two stages is required to maintain the proper reaction pH and to prevent the generation of any toxic cyanogen chloride (CNCl) gas, which forms at pH around 7.5 to 8 [7]. The final stage of the process involving cyanate oxidation is accomplished by lowering the pH of the solution after cyanate formation to 9, and addition of excess chlorine. Early operational difficulties associated with dispensing chlorine in high temperature solutions [1] have been overcome. Estimated reaction time for treating aqueous cyanide wastes with high-temperature alkaline chlorination is in the range of 2 to 3 h. For iron–cyanide-containing sludge materials, the treatment time can range from 6 to 9 h (unpublished data, Alcoa, Inc.). HTAC has been evaluated for the destruction of strong and weak metal–cyanide complexes associated with metal processing wastewaters, sludges, and spent potlining leachates [5–8]. A principal technical limitation of this technology lies in the incomplete destruction of cyanide because of the following possible reasons: (i) association of cyanide with other metal hydroxides that precipitate upon dissociation of metal–cyanide complexes; (ii) chlorine demand for metal oxidation; and (iii) presence of other constituents with high chlorine demand and different reaction kinetics. Complete destruction depends on reaction conditions including reaction time, pH, temperature, and chlorine dose. Reaction time can be critical if one is attempting to accomplish both cyanide and cyanate oxidation at the same pH. The selected reaction time has to be designed for CNO− oxidation, the rate limiting reaction. In addition, if other organics, such as, chlorinated aliphatics are present in the waste stream, use of additional oxidants may be necessary.

22.2.2 ACHIEVABLE TREATMENT LEVELS Total cyanide concentrations are typically reduced to a concentration less than 0.05 mg/l irrespective of the influent concentrations (unpublished data, Alcoa, Inc.).

22.2.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The controlling design parameters for high-temperature alkaline chlorination (HTAC) include temperature, reaction pH, and reaction time. Because of increased energy costs, this technology is usually suitable for treatment of small volumes of concentrated waste, for example, treatment flow rate less than 30 gpm. High-temperature alkaline chlorination is most suitable for treating waste streams with high metal–cyanide complex concentrations and the following characteristics: alkaline pH (pH > 9), TDS < 1000 ppm, and ORP: 500–550 mV. Optimal residence time is 2–3 h for aqueous streams, and 6–9 h for sludges at the operating temperature range of 165–180◦ F. Since it is customary to increase the pH of the waste (pH > 11 for liquid slurries and pH > 12 for sludges) prior to heating, pH adjustment is a key step in the HTAC process. Metal hydroxide sludges can be generated from the break-down of strong and weak metal complexed cyanides, which can interfere with cyanide treatment performance as explained earlier. Also, if aliphatic organics are present in the waste stream, chlorinated organics could be generated depending on reaction conditions.

22.2.4 COST OF THE TECHNOLOGY High-temperature alkaline chlorination is expensive to implement and operate. The capital costs for a full-scale 25 gpm continuous treatment system have been estimated (preliminary) at about $1.5M, and the operation and maintenance costs at about $0.42M/yr for a 25 gpm system (2001 cost basis, unpublished data, Alcoa, Inc.). For treatment of iron–cyanide-containing sludges, the capital and

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operations and maintenance costs for a 4500 gal/day system have been estimated (preliminary) as $1.4M and $0.6M/yr, respectively (2001 cost basis, unpublished data, Alcoa, Inc.).

22.2.5 TECHNOLOGY STATUS The HTAC technology has been implemented at bench-scale for treatment of clarifier sludge from an iron–cyanide, calcium–fluoride precipitation process at an Alcoa facility. There has been limited full-scale application of this technology at an Alcoa facility in Brazil for treatment of spent potlining leachate and leachate-impacted groundwater.

22.3 INCINERATION/THERMAL TREATMENT 22.3.1 PROCESS DESCRIPTION AND IMPLEMENTATION Incineration is employed for the thermal destruction of many waste materials. The process essentially entails the controlled combustion of the waste material at a desired temperature and is most suited for wastes that can sustain their own combustion (i.e., high in organic matter). Other wastes deficient in organic matter require supplemental fuel or blending with high organic wastes. Test studies have indicated high levels of cyanide destruction (>99.9%) during incineration of cyanide-containing waste materials [7]. Incineration has been commonly used to destroy cyanide wastes generated in organic chemical manufacturing, for example, acrylonitrile production. Thermal treatment has been effectively employed with cyanide-contaminated soils, sludges, leachates, wastewater, and spent potlining from aluminum smelting (unpublished data, Alcoa, Inc.). The basic components of thermal treatment include a primary and secondary combustion chamber followed by an air pollution control system, as shown in Figure 22.2. The residuals are fed to the primary chamber that typically operates at 1000 to 2000◦ F where the contaminants are volatilized into the gas phase and either pyrolized (in anoxic environments) or oxidized (in oxic environments)

Crushing

Receiving Storage

Dust recycle

Prepared potliner storage

Sand

Limestone

Gas cleaning

Kilns Landfill leachate

Off-spec recycle

Kiln product to landfill Off-spec recycle

FIGURE 22.2 Process flow diagram for Alcoa Gum Spring, Arkansas spent potliner thermal treatment facility. Courtesy of: Alcoa, Inc., Pittsburgh, PA, 2005.

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to combustion products like CO2 , H2 O, HCl, NOx , and SO2 . Carbon dioxide and H2 O are produced from carbonaceous matter, HCl from chlorinated organics, NOx from nitrogenous compounds including cyanide and nitrogen in the combustion air, and SO2 from sulfur compounds. This is a highly energy-intensive process depending on the energy value of the waste, and leaves only small volumes of residuals depending on the ash content of the waste. Thermal treatment is in commercial practice both as on-site and off-site operations. Off-site incineration is used for small volumes of soils, sludges, waste materials, and highly concentrated aqueous waste streams. To treat spent potlining material from aluminum smelting operations, various thermal treatment approaches have been tested, including combustion in a dry cement kiln (Alcoa Gum Springs, Arkansas), vitrification in a counter-current long rotary kiln (Nova Pb process), vitrification in a gas furnace (Alcoa Portland Ausmelt SPL process), vitrification in a Vortec furnace (Ormet process), and incineration in an electric arc furnace (SELCA process). Figure 22.2 shows the process flow sheet for Alcoa’s SPL thermal treatment facility at Gum Springs, Arkansas. Apart from use in the aluminum industry, the cement kiln process has also been applied for thermal destruction of cyanide-containing spent oxide box residuals from MGP sites [9]. All of these thermal processes destroy cyanide completely at the 1000 to 2000◦ F operating temperature, leaving no potential risk from vapor phase HCN emissions.

22.3.2 ACHIEVABLE TREATMENT LEVELS Incineration/thermal treatment can reduce total cyanide concentrations in wastes to nondetectable levels.

22.3.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The principal design factors for thermal treatment technologies are combustion temperature and chamber retention time. Thermal treatment is best suited for the treatment of heavily contaminated wastes with high organic content (high heating value). However, thermal treatment is sometimes applied to wastes with low heating values with the understanding that operating costs will be high because of the need for supplemental fuel or that other high organic wastes are available to off-set the need for supplemental fuel.

22.3.4 COST OF THE TECHNOLOGY No capital cost is involved if an off-site hazardous waste incineration/thermal treatment facility is used. Based on Alcoa’s Gum Spring, Arkansas operations, typical thermal treatment cost amounts to $330/ton for handling 70,000 tons of SPL material per year (1999 cost basis, unpublished data, Alcoa, Inc.). However, for construction of a new incineration/vitrification facility, capital costs could range anywhere from $25 to 50 million for a 100 ton/h process (2020 projected cost estimate, Alcoa, Inc.). This order-of-magnitude cost estimate is based on factored cost of major equipment components.

22.3.5 TECHNOLOGY STATUS Thermal treatment/incineration is a well-established, commercially practiced technology with significant public relations concerns. Commercial incinerators are available in the United States to treat hazardous waste, both as mobile as well as fixed-base operations. The fixed-base systems are generally located at offsite facilities and typically consist of large, rotary kilns ranging in size from 80 to 100 M BTU/h with typical solid capacities of 20 tons/h. The mobile units are primarily composed of various sizes of rotary kilns ranging between <20 M BTU/h to >40 M BTU/h with typical

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loading rates between 1 to 2 and >10 tons/h, or fluidized bed combustors with capacities about 4 tons/h.

22.4 ELECTROLYTIC DECOMPOSITION OR OXIDATION 22.4.1 PROCESS DESCRIPTION AND IMPLEMENTATION In this process, cyanide is oxidized electrochemically (anodically) under alkaline conditions. Concentrated cyanide waste composed of free and weak metal–cyanide complexes is subjected to electrolysis under alkaline conditions. Metals are separated from the complexes and deposited at the cathode and free cyanide is released into the solution. Free cyanide is thereafter oxidized at the anode with copper added to the waste solution as an electrochemical catalyst to further enhance the reaction rates. In addition, for dilute cyanide solution, addition of sodium chloride improves the decomposition process [10]. The anodic oxidation of free cyanide is proportional to the alkalinity of the waste solution. The final oxidation products are CO2 , N2 , and ammonia, with cyanate (CNO− ) formed as an intermediate. The entire reaction scheme is described by the following reactions [10]: CN− + 2OH− → CNO− + H2 O + 2e−

(22.5)

2CNO− + 4OH− → 2CO2 + 2N2 + 2H2 O + 6e−

(22.6)

2− CNO− + 2H2 O → NH+ 4 + CO3

(22.7)

In order to enhance the reaction rate, the process could be performed at higher temperatures (125 to 200◦ F) with ammonia produced as an additional by-product [11]. Anode materials include graphite, platinized titanium, lithium platinite, and nickel. Electrolytic decomposition is a well-established technique, primarily used by the electroplating industry for the destruction of high concentrations of free cyanide and weak metal–cyanide complexes in concentrated process waters. This technology is not suitable for strong metal–cyanide complexes, like iron and cobalt cyanides. Electrochemical oxidation becomes uneconomical at cyanide concentrations below several hundred ppm. In this case, conventional alkaline chlorination or other treatment procedures are used for final treatment [11].

22.4.2 ACHIEVABLE TREATMENT LEVELS Although electrochemical decomposition can reduce total cyanide levels to less than a few ppm, this technology has been shown to be cost effective only for reducing total cyanide levels of concentrated waste streams from >1000 mg/l (total cyanide) to less than a hundred mg/l (total cyanide) [11], at which point the waste stream is normally subjected to other, more efficient and effective technologies for treatment to discharge quality. Further reduction of cyanide is possible with electrochemical techniques if processing conditions involve sufficient reaction time (>6 h), conductivity (via NaCl addition), and oxidation potential [10,12].

22.4.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The critical design factors for the electrolytic decomposition process under ambient conditions are the current density (in amps/m2 ), volume of solution to be treated, and anode surface area. In general

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cyanide concentration is a linear function of the electrolysis time if the cyanide concentration is above 250 mg/l [10]. For a copper cyanide solution with a CN/Cu (mass) ratio of 4.6, Ho et al. [10] developed a design equation based on mass balance information and existing data as follows: C = C0 −

7 (0.25X + 4.9275) × (−0.369Y 2 + 12.945Y + 0.713)t V

(22.8)

where C0 and C are the initial and final total cyanide concentrations in mg/l; V is the volume of the solution (m3 ); X is the anode surface area (m2 ); Y is the applied current (amp); and t is the electrolysis time (h). The applicability of this equation is restricted to a particular CN/Cu ratio and the final cyanide concentration (>250 mg/l). For different compositions, however, the constants in the equation need to be modified. The theoretical energy requirement for this process is around 2.06 amp-h per kg of cyanide. At a cell voltage of 2 to 4 V, this would correspond to 4.1 to 8.2 kWh/kg of cyanide [13]. The electrochemical decomposition is an energy-intensive process and requires careful optimization of temperature (125 to 200◦ F) and pH (>10). The technology can be implemented as a batch or a continuous process. The electrochemical decomposition process is best suited for treating small volume process waters containing high concentrations of free and WAD cyanide, and metals of value [11]. The technology is most suitable for treating waste streams with high metal–cyanide complex concentrations and the following characteristics: pH > 11 and TDS < 1000 ppm. This technology is usually not suitable for treatment to low levels for discharge. High sulfate concentrations will cause heavy scaling at the anode and inhibits electrolysis. Hence periodic cleaning of electrodes is necessary when high sulfur concentrations are encountered in the process waters. Cyanate and ammonia are generated as residuals and could require alkaline chlorination treatment prior to discharge.

22.4.4 COST OF THE TECHNOLOGY Typical capital cost of electrochemical decomposition process has been reported as $50,000 for a 10 gpm system with operation and maintenance cost equivalent to $0.25/kg of cyanide destroyed (1985 cost basis [11]). This cost does not include the cost for treating any cyanide that is left untreated at the end of the operation.

22.4.5 TECHNOLOGY STATUS This electrochemical technology has been evaluated at bench- and pilot-scale for the treatment of high-strength cyanide wastes, and has been used in the electroplating industry. Different reactor electrode configurations have been tested, including parallel plate, packed bed, and bipolar configurations [12]. Overall, this technology is ready for implementation in the field for treatment of small-volume concentrated plating waste.

22.5 POLYSULFIDE PROCESS 22.5.1 PROCESS DESCRIPTION AND IMPLEMENTATION In this process, free cyanide and metal cyanides are reacted with an aqueous solution of ammonium polysulfide [(NH4 )2 Sx ] or calcium polysulfide [CaSx ] or nonalkyl ketones or aldehydes to form thiocyanate and other noncyanide end products. For the case where ammonium polysulfide is used

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as a reactant, the following reactions apply [14,15]: H2 O

HCN + (NH4 )2 Sx + NH3 −→ NH4 SCN + (NH4 )2 S(x−1)

(22.9)

where x is an integer; 2 ≤ x ≤ 5. H2 O

4+ (M)4+ [Fe(CN)6 ]4− + y(NH4 )2 Sx → FeS + 6SCN− + 2NH+ 4 + (y − 1)(NH4 )2 S2 + M (22.10) 4 where (M) is a metal ion or ions or NH+ 4 ; x = 3, 4, or 5; y = 4, 2, or 3 . This patented technology (Shell Oil Company) targets the treatment of strong and weak metal cyanides in purifier wastewater from gas scrubbing [14,15]. The process can be operated in batch or continuous mode with a typical residence time of 10 to 60 min at a temperature in the range of 110 to 180◦ C. The effluent stream following treatment may include iron sulfide (or other metal sulfide), thiocyanate, and metal hydroxides. In tests performed with spent potlining leachate and lime sulfur (a common source of calcium polysulfide), a CaSx /CN mass ratio of 4:1 was optimal for the formation of thiocyanate. This process is applicable for treating wastes with a flow rate between 50 and 100 gpm and cyanide concentrations in the range of 10 to 50 ppm [14]. Optimum process conditions are as follows:

• • • •

Temperature: 110 to 180◦ C Pressure: 1 to 10 atm pH: 8.5 to 9.5 Reactants: 0.01 to 1 M polysulfide solution (should be at least 3 to 4 times in excess of influent iron–cyanide complex concentration)

Any metal sulfide and hydroxide sludges generated should be removed from the effluent stream prior to subsequent biological treatment.

22.5.2 ACHIEVABLE TREATMENT LEVELS Baker [15] subjected 50 ml of aqueous slurry of fly ash particulates containing 150 mg/l of ferrocyanide complex to electrolytic decomposition with ammonium polysulfide in a closed vessel at 160◦ C for 30 min. Thiocyanate was produced and residual iron–cyanide complex concentration was below the detection limit of 10 ppb.

22.5.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The controlling design factors for the high-temperature polysulfide process are the reaction time, reaction pH, pressure, and the ratio between the cyanide and the polysulfide reagent [14,15]. This technology can be run in batch or continuous mode with typical residence time of between 10 and 60 min. It can be implemented in a single reactor and is relatively easy to operate. However, the operation requires constant monitoring of dosage, temperature, pressure, and pH. As far as residuals are concerned, metal sulfides, thiocyanate, ammonium ion, as well as metal hydroxide sludges can be generated. Thiocyanate production is minimal when treating spent potlining leachate using calcium polysulfide. Subsequent biological treatment of thiocyanate after removing all metal sulfides and hydroxide sludges from the effluent may be necessary prior to discharge. A similar but ambient-temperature polysulfide treatment process has been tested with concentrated electroplating wastewaters by Ganczarczyk et al. [16]. In tests with two electroplating

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wastewaters having 11,400 and 56,400 mg/l of total cyanide (consisting of weak metal–cyanide complexes and free cyanide), addition of calcium polysulfide (CN:polysulfide ratio of 1:1.5 on a w/w basis) reduced cyanide concentrations to levels less than 90 mg/l within one day at pH 10 to 11. Thiocyanate was produced in the process.

22.5.4 COST OF THE TECHNOLOGY No cost information for this technology was found in the open literature.

22.5.5 TECHNOLOGY STATUS In addition to multiple laboratory bench scale and pilot testing of this technology, sodium and ammonium polysulfide have been routinely used in catalytic cracking unit in refineries, where the polysulfide has been successful in converting over 90% of the cyanide ions to thiocyanate thereby preventing severe fouling and plugging [17]. For implementation of this technology, pilot-scale testing or at least some bench-scale testing is necessary on a case by case basis to develop site specific design criteria and operating conditions.

22.6 WET AIR OXIDATION 22.6.1 PROCESS DESCRIPTION AND IMPLEMENTATION The wet air oxidation (WAO) process entails oxidation of soluble and suspended organic waste components in an aqueous environment using oxygen as the oxidizing agent. The process utilizes moist air for oxygen and operates at elevated temperature (175–300◦ C) and pressure (20–200 atm) for treatment of municipal and industrial sludges. The process has also been used to treat wastes containing free cyanide and metal cyanides. The principal reactions involved in wet air oxidation of inorganic constituents like sulfides and cyanides are as follows: Cyanide species + O2 → CO2 + NH3

(22.11)

Sulfur species + O2 → SO2− 4

(22.12)

Organic N + O2

→

NH3 + CO2 + RCOOH

(22.13)

Patents on the WAO process have been issued to Zimpro Inc., Rothschild, Wisconsin [18,19]. Although these patents are not specific for cyanide oxidation, the process has been used to treat cyanide-containing sludges. Palmer et al. [7] reported on the operation of several full-scale Zimpro WAO systems used in treating cyanide-containing sludges. In these operations, total cyanides were reduced by 99.5%, with influent concentrations as high as 28,000 mg/l of total cyanide. This process is usually suitable for high strength wastewaters containing as high as 20,000 mg/l of cyanide that are difficult or uneconomical to treat via conventional biological treatment or thermal destruction (unpublished data, USFilter Corp.). The typical wet air oxidation system is a continuous process using high pressure rotating equipment to raise the feed stream and air (or oxygen) to required operating pressure [20]. Figure 22.3 presents a typical wet air oxidation process flow diagram. As shown in this figure, the waste is pumped through a high pressure slurry feed pump and the air is then combined with the slurry feed and passes through feed/effluent heat exchanger where the fluid is heated to the reaction temperature of about 300◦ C. The two-phase fluid enters the bubble reactor where the exothermic reactions (Equations [22.11] to [22.13]) take place. The residence time employed in the reactor is sufficiently long enough to enable the oxidation reactions to proceed to completion (∼1 h). WAO processing is

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Storage tank (by others) PC Feed effluent heat exchanger

Reactor

Off gas Cooling water out

L Process cooler

High pressure feed pump

M

P

Cooling water in

Pressure control valve

Treated effluent Process separator

Plant steam

Process air compressor

FIGURE 22.3 Wet air oxidation flow schematic. Courtesy of USFilter Corp. — Zimpro Products, Rothschild, WI.

energy intensive, and to conserve energy within the process the thermal energy in the treated effluent is recovered to preheat the incoming air-slurry mix entering the reactor as shown in Figure 22.3. Depending on the organic content or heating value of the waste, the exothermic nature of the reactions can produce sufficient energy in the reactor to allow the wet oxidation system to function without any additional heat input.

22.6.2 ACHIEVABLE TREATMENT LEVELS The WAO process has been applied to various waste streams containing cyanide, including military explosive waste, acrylonitrile, chemical plant wastewater, and other industrial wastewaters. Table 22.2 provides treatment performance data on various cyanide containing wastes that have been subjected to wet air oxidation. Residual total cyanide levels between 4 and 82 mg/l have been achieved for a wide variety of influent mass loading (123 to 29,800 mg/l of total cyanide) for the above mentioned waste matrices, while treated free cyanide level were as low as 0.1 mg/l for an influent free cyanide concentration of 870 mg/l.

22.6.3 TECHNOLOGY SELECTION FACTORS AND DESIGN CONSIDERATIONS The controlling design factors for WAO technology are reaction temperature and pressure. The process is usually performed in a pressurized hydrothermal reactor in the subcritical water temperature and pressure range. In order to maintain the influent feed in liquid form, it is necessary to maintain the system pressure above the steam pressure. The temperature range for cyanide destruction in industrial wastewaters is usually in the medium range between 175 and 220◦ C. Higher temperature systems (320 to 360◦ C) can be designed for complete destruction of organocyanide sludges but are rarely used due to high capital cost [20]. WAO technology can be applied to the treatment of complex waste streams containing COD ranging from 10,000 to 150,000 mg/l at a flow rate between 4 and 220 gpm [21]. Most often, a single reactor is employed to perform the complete oxidation step. The reactor is usually constructed of a wide variety of materials ranging from 316L stainless steel to titanium grade 12, capable of withstanding operating temperatures in the range of 400 to 600◦ F, pressures between 300 and 3,000 psig,

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TABLE 22.2 Treatment Performance Data for Wet Air Oxidation Type of cyanide species Total cyanide

Free Cyanide

Type of waste

Influent cyanide, ppm

Effluent cyanide, ppm

Military explosive waste Acrylonitrile Chemical plant wastewater Hazardous chemical wastewater Hazardous wastewater Acrylonitrile Chemical plant wastewater

123 1,946 21,040 25,340 29,800 870 20,030

4 1.5 44 82 9.5 0.1 61a

a At an intermediate temperature; an improved performance could be obtained by increas-

ing WAO temperature. Source: Data from Zimpro, USFilter Corp. — Zimpro Systems, Products, and Services, http://www.zimpro.usfilter.com.

and corrosive conditions. For smaller applications, skid-mounted units are available to treat flows of 20 gpm or less. Skid-mounted units are usually designed for intermediate range of operating conditions, with operating temperature ranging from 400 to 550◦ F and pressures from 750 to 2000 psig. Unlike other thermal processes, WAO produces no smoke, fly ash, or oxides of sulfur or nitrogen. Spent air from the system can be passed through an adsorption unit (e.g., carbon units) to meet local air quality standards.

22.6.4 COST OF THE TECHNOLOGY Detailed capital cost information for application of WAO for cyanide destruction depends on system size and is available from commercial vendors, for example, Zimpro [21]. Operating costs for running a WAO system for cyanide removal typically range from 2 to 6 cents/gallon (2001 cost basis, unpublished data, USFilter Corp.).

22.6.5 TECHNOLOGY STATUS The WAO process for hydrothermal destruction of cyanide has been a commercial technology since 1970. Over 200 industrial or municipal WAO systems have been constructed, many of which process cyanide-containing industrial wastewater and sludges. Large commercial units are available [21] that are capable of treating mixed cyanide waste streams with flow rates up to 220 gpm; smaller skid-mounted units are also available to treat waste streams with flow rates of 20 gpm or less.

22.7 OTHER THERMAL TECHNOLOGIES 22.7.1 THERMAL DESORPTION Thermal desorption uses heat to remove cyanide and organic contaminants, like PAHs, from surfaces of solid materials. Desorption occurs when the solid matrix is sufficiently heated causing the cyanide and organic fraction to volatilize. Effective temperature varies for different compounds. For cyanide-bearing solid wastes, desorption temperatures between 575 and 900◦ F have been effective [22].

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The effectiveness of full-scale thermal desorption was tested for manufactured gas plant (MGP) wastes by EPRI [22]. Almost 5000 tons of cyanide and PAH containing soils and sediments and tar emulsion wastes/leachates from MGP sites were treated in a gas-fired, co-current rotary kiln at temperatures between 600 and 1100◦ F and loading rates ranging from 11 to 23 tons/h to assess optimum conditions for cyanide, PAH, and BTEX destruction. Total cyanide concentrations in the residual were reduced from 7–700 to less than 3–30 mg/kg, a total decrease between 60 and 99%. EPRI [23] also tested electric utility boilers for thermal destruction of MGP solid wastes. Significant reductions in cyanide concentrations were observed with these techniques. The feasibility of using thermal desorption for treatment of aqueous cyanide wastes has not been evaluated. However, the temperature of these desorption processes are not sufficient to destroy the free cyanide in the offgas and therefore offgas treatment would likely be necessary. Usually, an afterburner is employed, where any offgases are incinerated.

22.7.2 THERMAL OXIDATION Thermal oxidation is another alternative technology for free cyanide and metal-complexed cyanide destruction involving either high temperature hydrolysis or combustion. It is very similar to HTAH technology, with the exception of the operating pH. While HTAH is performed strictly under alkaline conditions (pH > 12 units), thermal oxidation is usually operated at a neutral pH range. At temperatures between 140 and 200◦ C, pressures up to 100 bar, and a pH of 8, hydrolysis of cyanide occurs rapidly yielding formate and ammonia [13] according to the following reaction: CN− + 2H2 O → HCOO− + NH3

(22.14)

In the presence of nitrates, formate and ammonia can be destroyed in a tubular reactor at 150◦ C in the following reactions: − NH+ 4 + NO2 → N2 + 2H2 O

(22.15)

+ 2 3HCOOH + 2NO− 2 + 2H → N + 3CO2 + 4H2 O

(22.16)

The process can efficiently treat cyanide-containing waste streams over a wide concentration range and is applicable to both rinsewater and concentrated solutions.

22.7.3 THERMAL PLASMA Another treatment technology for the destruction of concentrated cyanide liquor involves direct contact with thermal plasma in a reactor [24]. The technology, which has been tested only at bench scale, consists of a single, direct-current plasma torch in a draft tube that is submerged in aqueous waste. Submergence of the plasma provides direct contact of the plasma and the contaminated water being treated. The thermal plasma technology has been successfully tested at bench scale for treatment of free and complexed cyanides in spent potlining leachate. Under atmospheric pressure and a reactor temperature of 100◦ C, the rate of cyanide decomposition by thermal plasma was about 12 times greater than that of thermal hydrolysis in a plug flow reactor at the same temperature [24].

22.8 THERMAL TECHNOLOGY OVERVIEW A summary of thermal technologies for treatment of solids, sludges, and wastewater containing cyanide is given in Table 22.3. This table summarizes the key technology features that have been described, and can be used to assist technology screening efforts.

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Free CN

X

X

X

Technology

Polysulfide process

High temperature alkaline hydrolysis

High temperature alkaline chlorination X

X

X

WAD CN

Chemical applicability

X

X

X

FeCN

TABLE 22.3 Thermal Technology Screening Matrix

This technology involves the same underlying principle used in the alkaline chlorination process, except that it uses a high temperature of approximately 170◦ F to destroy FeCN. Estimated reaction time for treating aqueous influents range from 2 to 3 h. For FeCN sludge the treatment time is between 6 and 9 h

This technology involves the treatment of FeCN sludges and leachates through hydrolytic cracking at high temperature and pressure under alkaline conditions. The liberated free CN is converted to formate and ammonia at elevated temperatures. Formate and ammonia are further oxidized to H2 O, CO2 , and N2

This patented technology by Shell Oil Company involves the treatment of CN-bearing purifier wastewater by contacting it with an aqueous solution of ammonium polysulfide, (NH4 )2 Sx , calcium polysulfide (CaSx ) or nonalkyl ketones of aldehydes. The process may be run in batch or continuous mode with a typical residence time of 10–60 min within a temperature range of 110–180◦ C and an optimal CaSx /CN mass ratio of 4:1

General description Not available

$2.2M for a 25 gpm GW system

$1.5M for a 25 gpm GW system

<0.1 mg/l total CN and <4 mg/l soluble Fe, irrespective of the influent CN conc.

<0.05 mg/l total CN for SPL leachate/GW, influent >50 mg/l total CN

Capital

0 to 5 mg/l for a FeCN influent of 150 mg/l

Achieveble treatment levels

$0.42M/yr for a 25 gpm GW system

$0.44M/yr for a 25 gpm GW system

Not available

O&M

Costs

∼100 K/yr for off-site transport and nonhazardous landfill disposal, for 25 gpm system

Minimal

Not available

Waste mgmt.

Implemented at bench scale for treatment of SPL leachate and clarifier sludge from iron salt and fluoride precipitation sludge

Implemented at pilot-scale at an Alcoa facility for SPL leachate treatment of cyanide

Limited field-scale implementation. Only 2 to 3 actual field applications documented

Technology status

454 Cyanide in Water and Soil

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X

X

Wet air oxidation

X

Electrolytic decomposition

Incineration/ thermal treatment

X

X

X

X

X

Wet air oxidation involves oxidation of soluble or suspended components in an aqueous environment using oxygen as the oxidizing agent. The wet air oxidation (WAO) process utilizes moist air at elevated temperature (175–300◦ C) and pressure (20–200 atm) to oxidize mixed cyanide wastes from municipal and industrial sludges, containing free CN as well as metal–cyanide complexes.

Electrolytic decomposition is primarily used by industry for the destruction of CN in concentrated electroplating wastes. The technology can treat waste streams contaning high concentrations of free and WAD CN. The concentrated CN waste is subjected to anodic electrolysis under alkaline conditions, whereby the metal is seperated from the complex at the cathode and the free cyanide is released into the solution, which is subsequently oxidized to CO2 , N2 , and ammonia, with cyanate (CNO− ) as an intermediate

This technology involves complete burning of cyanide containing residuals (solids and leachates). This is a conventional incineration process and leaves practically no residual by-products

2 to 6 cents per gallon of waste treated

Available from USFilter Zimpro products

4 to 82 mg/l total CN for a total CN influent ranging from 123 to 30,000 mg/l; Free CN ∼0.1 mg/l for influent ∼870 mg/l

$0.25/kg CN destroyed

∼ $50K for a 10 gpm system

<0.4 mg/l total CN for a total CN influent >50, 000 mg/l

$330/ton incineration cost

No capital costs: assumes incineration facility in place

Nondetectable cyanide in the effluent

Minimal

Minimal

Minimal

Large commercial and small skid-mounted units are available

Demonstrated at bench- and pilot-scale. One small-scale field application also documented

Commercial incinerators are available in the U.S. to treat hazardous waste

Thermal and High Temperature Oxidation Technologies

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22.9 SUMMARY AND CONCLUSIONS • Most thermal treatment technologies are capable of destroying free cyanide as well as weak and strong metal–cyanide complexes to benign end-products, like CO2 , N2 , or NH3 . Electrolytic decomposition technology is suitable only for treatment of free cyanide and weak metal–cyanide complexes. • Most thermal technologies involve energy-intensive processes and require high temperature to break down the contaminants in the waste stream. They can reduce waste volumes significantly, leaving only a relatively small amount of hazardous waste residue. Some technologies, like the electrolytic decomposition and polysulfide processes, however, produce certain by-products, like ammonia and thiocyanate, that could require additional treatment prior to discharge. • Thermal technologies capable of destroying strong metal–cyanide complexes are expensive to operate and are considered economical only for treatment of small volumes of high concentration wastes. • Some thermal technologies, including high-temperature alkaline hydrolysis, incineration/ thermal treatment, and wet air oxidation have been field demonstrated for soils, sludges, and solid and aqueous wastes containing cyanide.

REFERENCES 1. Robuck, S.J. and Luthy, R.G., Destruction of iron-complexed cyanide by alkaline hydrolysis, Wat. Sci. Tech., 21, 547, 1989. 2. Kimmerle, F.W., Girard, P.W., Roussel, R., and Tellier, J.G., Cyanide destruction in spent potlining, in Proceedings of Light Metals 1989, The Minerals, Metals and Materials Society, Warrendale, PA, 1989. 3. Futakawa, M., Takahashi, H., Inoue, G., and Fujioka, T., Treatment of concentrated cyanide wastewater, Desalinization, 98, 345, 1994. 4. Bratina, J., Heritage Environmental, Inc., personal communication, 2004. 5. Blayden, L.C., Hohman, S.C., and Robuck, S.J., Spent potliner leaching and leachate treatment, in Proceedings of Light Metals 1987, The Minerals, Metals and Materials Society, Warrendale, PA, 1987, p. 663. 6. Wedl, D.J. and Fulk, R.J., Cyanide destruction in plating sludges, Metal Finish., 89, January 33, 1991. 7. Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Surprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Noyes Data Corp., Park Ridge, NJ, 1988. 8. Hassan, S.Q., Vitello, M.P., Kupferle, M.J., and Grosse, D.W., Treatment technology evaluation for aqueous metal and cyanide bearing hazardous wastes (F007), J. Air Waste Manage. Assoc., 41, 710, 1991. 9. Hayes, T.D., Linz, D.G., Nakles, D.V., and Leuschner, A.P., Management of Manufactured Gas Plant Sites, Vol. II, Amherst Scientific Publishers, Amherst, MA, 1996. 10. Ho, S.P., Wang, Y.Y., and Wan, C.C., Decomposition of cyanide effluent with an electrochemical reactor packed with stainless steel fiber, Water Res., 24, 1317, 1990. 11. Patterson, J.W., Cyanide, in Industrial Wastewater Treatment Technology, 2nd ed., Butterworth, Boston, MA, 1985, p. 115. 12. Lin, M.L., Wang, Y.Y., and Wan, C.C., A comparative study of electrochemical reactor configurations for the deposition of copper cyanide effluent, J. Appl. Electrochem., 22, 1197, 1992. 13. Hartinger, L., Handbook of Effluent Treatment and Recycling for the Metal Finishing Industry, 2nd ed., Finishing Publications, Ltd., Stevenage, Herts, U.K., 1994. 14. Baker, D.C., Process for the removal of iron cyanide complex or complexes from aqueous solution, U.S. Patent 4,654,148, 1987. 15. Baker, D.C., HCN and iron cyanide complex removal, U.S. Patent 4,693,873, 1987. 16. Ganczarczyk, J.J., Takoaka, P.T., and Ohashi, D.A., Application of polysulfide for pretreatment of spent cyanide liquors, J. Water Poll. Control Fed., 57, 1089, 1985. 17. Kunz, R.G., Casey, J.P., and Huff, J.E., Refinery cyanides: a regulatory dilemma, Hydrocarbon Process., October, 98, 1978.

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18. Schoeffel, E.W. and Zimmerman, F.J., Wet air oxidation of combustible materials adsorbed on carbonaceous adsorbent, U.S. Patent 3,386,922, 1968. 19. Zimmerman, F.J., Waste disposal, U.S. Patent 2,665,249, 1950. 20. Maugans, C.B. and Ellis, C. Wet air oxidation: a review of commercial subcritical hydrothermal treatment, in Proceedings of IT3’02 Conference, New Orleans, LA, 2002. 21. Zimpro, Zimpro wet air oxidation system, USFilter Corp. — Zimpro Systems, Products, and Services, http://www.zimpro.usfilter.com, accessed: April 25, 2005. 22. EPRI, Field test of manufactured gas plant remediation technologies: thermal desorption, TR-105145, Electric Power Research Institute, Palo Alto, CA, 1995. 23. EPRI, Remediation strategies for source materials and contaminated media at manufactured gas plant (MGP) sites, TR-103811, Electric Power Research Institute, Palo Alto, CA, 1994. 24. Soucy, G., Fortin, L., Kasireddy, V.K., Bernier, J.L., and Kimmerle, F.M., Innovative cyanide solution treatment by thermal plasma, in Cyanide: Social, Industrial, and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals, and Materials Society, Warrendale, PA, 2001, p. 379.

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Technologies 23 Microbiological for Treatment of Cyanide George M. Wong-Chong and Jeanne M. VanBriesen CONTENTS 23.1

Process Fundamentals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.1 Inherent Biodegradability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.2 Presence of Capable Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.3 Availability of Critical Nutrients . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.4 Absence of Inhibitory Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.5 System Conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.6 Contact Time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.1.7 Competitive Processes in Biotreatment Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2 Wastewater Treatment Process Configurations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.1 Suspended Growth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.1.1 Batch, SemiBatch, and Sequencing Batch Reactors . . . . . . . . . . . . . . . . 23.2.1.2 Activated Sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.2 Attached Growth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.2.1 Trickling Filters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.2.2 Rotating Biological Contactors. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.2.3 Fluidized Bed Reactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.3 Novel Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.4 Anaerobic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.2.5 Process Limitations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.3 Full-Scale Applications of Biological Treatment for Cyanide in Gold Mining . . . . . . . . . . 23.4 Biological Treatment of Solid Wastes Containing Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

460 460 461 462 462 463 463 463 464 464 464 465 468 468 468 469 470 471 471 471 472 474 474

Microbiological systems for the treatment of wastewaters were developed around the late 1890s, initially for treatment of human waste and later for treatment of industrial wastes. The first treatment systems were run in “batch mode”; wastewater was held in reactors containing rocks. Although the microbial nature of the process was not clearly understood initially, it is now known that microbes grow in biofilms attach to the rocks, utilize carbon and nitrogen compounds in the waste as a food source, and thus reduce the oxygen demand the wastewaters would exert when released to the environment. To increase the amount of wastewater that could be treated and decrease problems associated with clogging, two continuous processes were developed in the early 1900s — trickling filters [1] and activated sludge [2]. These processes continue to dominate wastewater treatment systems to the modern day. Expanding chemical production in the 20th century led to increases in the number of anthropogenic compounds in sanitary and industrial waste streams that required chemical or biological 459

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treatment. Pretreatment or treatment of industrial waste streams often utilizes conceptually similar types of systems as conventional sanitary treatment; however, process parameters are usually optimized for specific target compounds. Biological treatment systems today are utilized to treat many anthropogenic compounds including aromatic hydrocarbons, halogenated solvents, BTEX compounds, and cyanide. Microbiological processes serve as the basis for robust treatment technologies because the removal of the contaminant enhances growth of the microbial biomass, and the increased biomass results in an increased rate of removal of the contaminant. This autocatalytic characteristic allows biological processes to remove substrate concentrations to very low levels. Many compounds can be completely transformed into innocuous compounds (e.g., CO2 and H2 O), and thus, biological processes generally produce only biomass (and sometimes gases, e.g., methane) as a waste product requiring further management. When the target compound can be used as the electron donor or carbon source for cell synthesis and energy generation, biological treatment systems are highly successful. Targeting the electron acceptor can also be successful (e.g., denitrification or reductive dechlorination), but this may require more careful attention to process dynamics. Suspended and attached growth systems in numerous process configurations are routinely used for domestic and industrial waste streams (see extensive examples in Rittmann and McCarty [3]). Despite the widespread economic success of biological treatment for many industrial waste streams and the use of biological treatment for some types of cyanide wastes for many years, the treatment of cyanide-bearing wastewaters by biological processes has not been extensively exploited for gold mining and other wastewaters in which cyanide species are present in relatively large concentrations and are primary treatment targets. The toxicity of free cyanide and weakly complexed metal cyanides (weak acid dissociable (WAD), cyanide), as well as a lack of understanding of the fundamental microbiological processes, possibly has slowed the commercialization of biological treatment of cyanide. Specifically, the critical role of chemical speciation in controlling biodegradation, while long known, has been a challenge to the design and operation of successful cyanide biotreatment systems. Despite the challenges, biological processes have the potential for treatment of a broad range of cyanide-bearing wastewaters, reducing cyanide to very low concentrations. This chapter provides an overview of the microbiological treatment of cyanide-containing wastewaters. Fundamental controlling parameters and system configurations with commercial applications are discussed.

23.1 PROCESS FUNDAMENTALS Effective biological treatment of contaminated waters depends on the biodegradability of the contaminants as well as on critical process parameters. The latter include (1) presence of capable microorganisms; (2) availability of carbon source, nitrogen source, electron donor, electron acceptor, and any essential trace nutrients; (3) absence of inhibitory materials; (4) system conditions including pH, temperature, and salinity; and (5) contact time between the microorganisms and the target pollutant(s).

23.1.1 INHERENT BIODEGRADABILITY Free cyanide (HCN, CN− ) is biodegradable under aerobic and anaerobic conditions. It can be used as the sole nitrogen or the sole carbon and nitrogen source for the growth of microbial systems. The utilization of different forms of cyanide (e.g., free cyanide, thiocyanate, and metal-complexed cyanide) follows different pathways and may depend upon other system conditions. Chapter 6 provides details about the biodegradability of cyanide. The dominant mechanism used in biological treatment systems for cyanide that have been scaled up to commercial applications appears to be oxidation to carbon dioxide and ammonia. The formed

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ammonia is then used as nitrogen source and incorporated into the cell material or can be utilized as an electron-donor substrate for obligate aerobic nitrifiers. Uncomplexed cyanide as well as thiocyanate can be transformed via the oxidative pathway. Equations (23.1) and (23.2) are typical cyanide degradation reaction equations: HCN + O2 + H+ + NADH → CO2 + NH3 + NAD+

(23.1)

+ SCN− + 3H2 O + 2O2 → CNO− + HS− + 2O2 → SO2− 4 +H

(23.2)

CNO− + 3H+ + HCO− 3

→

NH+ 4

+ 2CO2

A more comprehensive list of reactions involved with aerobic biodegradation of cyanide species is presented in Table 6.2. The effect of speciation on biodegradation of cyanide has long been known [4]. Current best interpretation of the available microbiological studies and wastewater treatment reports suggests that effective biological degradation of cyanide is limited to free and WAD cyanide. While wastewaters containing strong-acid dissociable (SAD) and iron–cyanide complexes may be degradable through dissociation (either acid/pH or UV-light induced), degradation rates in these systems tend to be too slow to be of commercial value.

23.1.2 PRESENCE OF CAPABLE MICROORGANISMS The microbiological studies summarized in Chapter 6 identified a wide variety of microorganisms (bacteria, fungi, and yeast) with inherent or inducible capabilities to degrade uncomplexed cyanide (HCN, CN− ) and certain metal-complexed cyanides (see Table 6.1). Many common species are represented, including arthrobacter, bacillus, and pseudomonas species. These groups are likely to be present in the mixed heterotrophic population found in wastewater treatment plants. Cyanide-degrading organisms often exist together with nitrifying bacteria. Nitrifiers are obligate aerobes, utilizing ammonia as an electron donor and oxygen as terminal electron acceptor and co-substrate. Nitrifying organisms have been extensively studied, and their growth rates and sensitivities are well documented. As chemoautotrophs utilize inorganic carbon as carbon source, they have low yields. An upset in cyanide transformation will not only reduce the available ammonia for the nitrifiers, but will also expose them to levels of cyanide that may inhibit their respiratory chain [5]. The slow growth of nitrifiers can make recovery of system performance difficult following an upset [6]. In current practice, all biological treatment systems operate with mixed microbial populations, and design of biological treatment is generally predicated on control of the slowest growing organisms. In systems where nitrifiers are utilizing the ammonia formed during cyanide degradation as electron donor, these organisms are typically growing much slower than the cyanide-transforming organisms. [7] Thus, control of a system designed to remove cyanide from a wastewater and to discharge it into waters with limits on ammonia may have to be designed for conditions that are controlled by the growth of nitrifiers rather than cyanide-degraders. Consequently, these systems tend to operate with long cell-retention times to avoid washout of the slow-growing nitrifiers. While biological treatment processes are dominated by whole-cell, growth-based systems, utilization of immobilized enzymes rather than whole cells is sometimes considered when the target compound is not a primary growth substrate. This technique has been considered for cyanide biotransformation. Two companies currently market immobilized enzyme systems. Imperial Chemical Industries in the U.K. markets a product called “Cyclear” [8,9] and Novo-Nordisk A/S in Denmark markets a cyanidase product [10]. Cyclear is based on extruded and freeze-dried fungal mycelia. These products work well for free-cyanide-containing wastes but are ineffective with complexed cyanides. While no commercial-scale operation exists, this method may be economically viable for treating very small quantities of wastewater.

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23.1.3 AVAILABILITY OF CRITICAL NUTRIENTS The culturing media for organisms described in Chapter 6 were generally rich in organic matter as a carbon source for microbial growth (e.g., broth, glucose, etc.) and the cyanide served as nitrogen source [7,11]. Because the organisms transform cyanide in order to utilize the nitrogen, there is potential for the availability of other nitrogen sources to interfere with cyanide removal. Dursun et al. [12] report inhibition of complexed cyanide removal when alternative nitrogen sources were available, and Mudder [13] indicates that the potential for alternative nitrogen sources to inhibit cyanide transformation should be considered during process design. Conditions can also be maintained that encourage utilization of cyanide as both nitrogen and carbon source; however, these are unlikely in conventional wastewater treatment systems where organic compounds are typically abundant. Alternatively, if carbon and nitrogen are supplied, some organisms will still transform cyanide as a detoxification mechanism [14]. This type of fortuitous degradation is not generally exploited in biological process design, where contaminants are targeted as growth substrates in order to take advantage of the autocatalytic nature of microbial degradation. Detoxification of cyanide in the absence of anabolic utilization of the released carbon or nitrogen will not be discussed further here. While anaerobic cyanide biotransformation does not require oxygen (see discussion of anaerobic biotransformation in Chapter 6), the oxidative pathway is the most exploited in biological treatment of cyanide-containing wastewaters. This pathway relies on oxygenase-enzymes, which require oxygen as a co-substrate in addition to its use as a terminal electron acceptor. This fact, in addition to the requirement for oxygen as an electron acceptor in nitrification, means that biotreatment of cyanide-containing wastes is most often designed to operate as an aerobic system.

23.1.4 ABSENCE OF INHIBITORY COMPOUNDS While the susceptibility of cyanide-degrading organisms to cyanide toxicity is not expected to be high, nitrifiers show high sensitivity to cyanide [15–17]. In the absence of nitrifying bacteria, ammonia can accumulate during cyanide biodegradation. Ammonia is known to be inhibitory to microbial processes, especially nitrification processes, at high concentrations [18,19]. However, in commercial scale treatment of coke plant wastewater where ammonia generally is severalfold greater in concentration than cyanide, inhibitory effects of ammonia are not usually observed under normal operating conditions [20,21]. However, under upset conditions inhibitory effects could prevail. For example, in a pilot-plant study of the effects of ammonia shock loading on a nitrification process used to treat petrochemical wastewater [22], total upset of a nitrification system was observed when the influent ammonia concentration was raised from 110 to 320 mg/l in 1 day. Since cyanide exists in many complexed forms (see Chapter 5), one potential result of WAD cyanide biodegradation is the release of previously complexed metals into solution. While many heavy metals (i.e., Cd, Cr, Cu, Pb, Ni, Zn, As, Hg, and Ag) can exert toxic effects on microbial activity, copper and zinc have effects at the lowest concentrations [23]. Whitlock and Smith [24], and Akcil and Mudder [6] reported the adsorption and accumulation of metals in their treatment system biomass. Accumulation of toxic/inhibitory metals (e.g., copper) in the biomass could account for the poor performance of early activated sludge tests. During the biological treatment of WAD cyanide (e.g., copper and zinc), metal hydroxides are formed. Those metal hydroxides can react with the strong iron–cyanide complexes to form a bi-metallic, solid-phase complex according to the following equation: 2Cu2+ + Fe(CN)−4 6 → Cu2 Fe(CN)6 (s)

(23.3)

The copper–iron–cyanide complex is a low-solubility precipitate that will settle out in the biological treatment tank or agglomerate with the biomass. The biological mass in the treatment system can

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secrete complex polymers, which tend to scavenge the precipitated metal complexes from the solution. The net result is the removal of metallic ions from the solution and, potentially, metal-complexed cyanide removal.

23.1.5 SYSTEM CONDITIONS Culturing conditions for organisms described in Chapter 6 were ambient mesophilic temperatures (30◦ C optimal). With commercial industrial wastewater treatment systems it is desirable to operate at ambient temperature conditions where temperatures can range from low levels during winter to 30◦ C plus in summer (e.g., gold mining operations at Homestake, Lead, South Dakota, operate at 10 to 30◦ C) [24]. Other systems (e.g., coke plant treatment systems) are more temperature sensitive and require mesophilic temperatures (25 to 35◦ C). pH conditions are generally in the neutral range. However, halophilic organisms with cyanide-utilizing capabilities at pH conditions about 10.0 have been isolated, [25] and some organisms have their optimal pH range well above neutral (e.g., 9.2 to 11.4) [7]. Coke plant treatment systems operate at pH in the range of 6.7 to 7.5 [17,26], and the Homestake gold mining tailings treatment system operates with feed water in the pH range of 7.5 to 8.5 [24]. Cyanide chemistry, as described in Chapter 5, includes significant aqueous complexation reactions that lead to multiple forms of cyanide in solution under typical wastewater conditions. Based on the concentrations of other species in the water, cyanide can appear in multiple forms. As discussed in Chapter 6 and briefly above, free cyanide is known to be amenable to biotransformation [27]. In most studies of complexed cyanide degradation, it was assumed that the complexed cyanide was present in equilibrium with free cyanide, and it was the free cyanide species that were utilized by the microorganisms. The relative concentration of complexed and uncomplexed cyanide in the system depends on other chemicals that are present and on the stability of the complexes that form. The importance of predicting the concentration of the bioavailable substrate form in experimental systems in order to evaluate bioremoval potential and transformation rates for chelates was described by VanBriesen et al. [28]. Cyanide has been observed to be removed even when the concentration of free cyanide is very low [27,29]. The presence of Ni2+ , Co2+ , Mn2+ , and Mo2+ in solution with cyanide reduced the degradation rate [30]. Further, Dumestre et al. [31] calculated species-specific rate constants (kHCN = 0.0714 h−1 and kCN = 0.0017 h−1 ). They also suggested that the often observed pH dependence for the rate of cyanide transformation is related to the decreasing concentration of the preferred substrate form (HCN) rather than to any decrease in enzyme action at high pH.

23.1.6 CONTACT TIME Biological systems must be designed to allow sufficient time for the active microorganisms to be in contact with the target compounds. This contact time is achieved by selecting a mean cell retention time (MCRT) or solids retention time (SRT) and adequate hydraulic retention time (HRT). MCRT or SRT assures maintenance of the desired microbial density and HRT is the actual contact time between wastewater contaminants and the microbes. As mentioned above, nitrifiers grow more slowly than cyanide-transforming organisms and thus SRT is selected to avoid washout of the nitrifiers. For conventional municipal treatment systems, this requires an SRT of 12 to 15 days. However, for industrial wastewaters like those from coke plants, the required SRTs can be in the range of 40 to 100 days.

23.1.7 COMPETITIVE PROCESSES IN BIOTREATMENT SYSTEMS In the biotreatment of cyanide there are concerns regarding competitive processes with the potential to remove the contaminant from aqueous solution, thus reducing its availability to the organisms.

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There is the perception that cyanide air stripping may occur in aerobic treatment systems; however, for properly developed systems, complete removal by biotransformation was achieved and air stripping was not significant [15,16,32]. Whitlock and Smith [24] reported significant adsorption of metal-cyanide species to biomass, but it is unclear if this enhances bioavailability or sequesters cyanide from further transformation.

23.2 WASTEWATER TREATMENT PROCESS CONFIGURATIONS Microbiological treatment of wastewaters can be accomplished in a variety of processes that are generally divided into those that are based on suspended growth of organisms and those that are based on the use of microorganisms growing attached to surfaces in a biofilm.

23.2.1 SUSPENDED GROWTH Suspended growth processes involve bacteria suspended in aqueous solutions that contain the substrates and nutrients needed for their growth. Organisms aggregate into flocs and utilize substrates directly from the bulk media with little or no mass transfer limitations. Suspended growth processes can be operated for a wide variety of applications including aerobic and anaerobic treatment of biological oxygen demand (BOD), nitrification, and denitrification. Suspended growth processes are generally divided into those with once-through treatment and those with recycle of active solids. 23.2.1.1 Batch, SemiBatch, and Sequencing Batch Reactors A batch reactor system is a large reactor that is operated in nonsteady-state mode. Microbial mass (in the form of sludge from a previous batch), the wastewater, and any essential nutrients are added to the reactor and allowed to react. Over time, the microorganisms grow, utilize the substrates in the waste, and settle at the bottom of the reactor. Periodically, sludge is removed from the reactor. Aeration and mixing may be utilized, depending upon the system. If the initial sludge loading is low, (i.e., low microbial population) long detention times are often needed for the organisms to reduce high waste concentrations to low concentrations. Lagoons are the most common example of this type of process; detention times of months to years are not unusual, and oxygen transfer depends upon wind action over a large surface area. Batch processes are also common for anaerobic processes where slow-growing organisms do not require frequent removal of biomass from reactors. A sequencing batch reactor (SBR) is essentially a batch reactor operated in fill-and-draw mode. SBR systems generally comprise two vessels, with each vessel operating as both reactor and clarifier in a sequence of five steps as shown in Figure 23.1: (1) fill, (2) reaction/aeration, (3) setting/clarification, (4) decant, and (5) idle. These steps vary in duration depending on performance requirements and wastewater characteristics. A two-vessel system would operate in series; one vessel filling and reacting while the other is in settling stage. Boucabeille et al. [33] demonstrated the removal of cyanide in batch, fed-batch, and continuous systems using a bacterial seed developed from a cyanide wastewater storage basin soil. Removal of cyanide was accompanied by formation of ammonia, and nitrification was observed either concurrently or following cyanide removal. Possibly due to concerns regarding the potential for substrate toxicity/inhibition during the initial fill stage, there is no indication in the literature that traditional suspended growth SBRs have been evaluated for their potential for cyanide biotreatment. For example, the treatment of coke plant wastewaters (described below) likely would require very large reaction basins to counteract the potential inhibitory effects of cyanide on thiocyanate degradation and ammonia nitrification. In the treatment of mining and electroplating wastewaters, there would be the additional concern for heavy

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Influent

Purpose

Operation

Typical time (h)

Add wastewater

Aeration on or off

1.5–3

Reaction time

Aeration on

1.5–3

Clarification

Aeration off

0.75–1.5

Withdraw effluent

Aeration off

0.5–1

Cycle complete

Aeration on or off

0.25–0.5

Recirculated Mixed Liquor

Effluent

FIGURE 23.1 Sequencing Batch Reactor Operating Sequence, Courtesy of U.S. Filter Corp.

metals inhibition. A sequencing batch biofilm reactor (SBBR) operating with a 48-h cycle time was tested for treatment of leachate from spent ore from gold mining [34]. The pilot system was able to treat 0.5 mg cyanide/l-h. 23.2.1.2 Activated Sludge Figure 23.2 presents a schematic illustration of a typical activated sludge process. As shown, it comprises a reaction basin in which microorganisms are maintained in suspension by either mixing or aeration, or both, to allow contact between substrate constituents and the microorganisms. The mixture of microorganisms and wastewater (the mixed liquor) flows to a clarifier where the solids are separated from the treated wastewater. The separated sludge is removed in the clarifier underflow, a major portion is recycled to the reaction basin to contact new incoming wastewater, while the remainder is wasted and goes to disposal. The return of organisms to the aeration basin allows for

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Clarifier (sedimentation) Effluent Wastewater

Aeration tank

Air or oxygen Return activated sludge

Waste activated sludge

FIGURE 23.2 Schematic flow diagram of conventional activated sludge process.

the maintenance of a predetermined concentration of microorganisms in the aeration basin. The rate of wasting of the separated biosolid sludge allows for the maintenance of the “sludge age” or MCRT or SRT. Since the water moves through the system only once and the organisms are recycled, SRTs are greater than HRTs. Thus, a smaller tank with long SRT can be used, and still achieve reasonable conversion rates for the contaminants in the wastewater. Typical MCRTs in conventional municipal wastewater treatment are in the order of 4 to 15 days, with longer times being characterized as “extended aeration” and generally used to encourage nitrification. Typical hydraulic residence times are in the order of hours rather than days. Industrial treatment systems can operate with SRTs as high as 100 days and HRTs in the order of 1 to 4 days depending on the concentrations of contaminants in the wastewaters. For activated sludge systems that treat municipal waste, the carbonaceous and sometimes nitrogenous compounds in the influent are used as electron-donor substrates by the suspended organisms, while injected air provides oxygen as an electron acceptor and mixing. However, other configurations are possible such as nitrification/denitrification systems where multiple tanks are used under different conditions to allow for use of ammonia and carbonaceous materials as electron donors, and oxygen and nitrate as acceptors in order to facilitate conversion of carbon and nitrogen to CO2 , N2 , and cell biomass. Tertiary denitrification systems generally require the use of an external carbon source (e.g., methanol), while predenitrification systems use the inherent carbonaceous content of the wastewater as electron donor but require recycling of nitrified mixed liquor. Cyanide removal has been reported in aerobic activated sludge systems by several groups [35–38]. Biological treatment in activated sludge plants has been reported to degrade up to 200 mg/l cyanide [39,40]. Neufeld and colleagues [41] evaluated the effect of process parameters for activated sludge systems that treat thiocyanate containing waste streams. Sludge ages less than six days exhibited process instabilities. Acceptable performance was observed with pH values of 5 to 7, indicating process insensitivity in near-neutral pH. Sludge settling in the pure culture reactors was excellent. Desai and Ramakrishna [42] report on the development, by the Indian Petrochemicals Corporation Ltd., of an extended aeration-activated sludge process as a second stage in the treatment of 300 m3 /day of cyanide-containing wastewater. Figure 23.3 illustrates the degradation profile for coke plant wastewater constituents in activated sludge treatment [15–17]. These degradation profiles show the following: • Organic matter represented by phenol was easily and readily degraded even in the presence of high concentrations of cyanide, thiocyanate, and ammonia, and degradation followed a linear, zero-order reaction profile.

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120 (7.1)

250

(7.2)

NH4+

100

200

ØOH

80

150

70 100 SCN–

40

NH3 concentration, mg/l

Concentration, mg/l

(7.2)

(7.2)

467

50 20 CN 0

0

2

4

6

8

10

12

14

16

Reaction time, h

FIGURE 23.3 Degradation profile for coke plant wastewater constituents. (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute to the American Iron and Steel Institute 1976. With permission.)

• Cyanide was removed after phenol, and degradation followed an exponential first-order reaction profile. With cyanide degradation there was a release of ammonia into the medium. • Thiocyanate was removed when the free cyanide concentration was reduced to less than 5 mg/l. Degradation followed a linear zero-order reaction profile, and ammonia was also released. • Ammonium–nitrogen is the last parameter to undergo degradation and this occurred when the cyanide concentration was less than 0.5 and thiocyanate concentration less than 3 mg/l. • With discharge limits on ammonia–nitrogen at less than 15 mg/l, ammonium degradation becomes the controlling reaction. The treatment rate developed for this parameter must reflect the inhibitory influences of cyanide and thiocyanate; these rates can be developed by bench- or pilot-scale treatability testing. Additional details on the treatment of coke plant wastewaters are presented in Chapter 26. Despite these promising results, questions have arisen regarding the fate of cyanide in aerobic biological treatment processes. Stripping during aeration was found responsible for 15% of the total cyanide removed in aerobic treatment units [37]. Another 12 to 16% was reported as removed through entrainment with the flocculated and settled cells [37]; however, others have observed no transfer to the sludge [39]. Cyanide in sludge was found to be predominately complexed and unavailable for subsequent degradation. In studies of coke plant wastewater treatment with properly developed and acclimated sludge, no losses of cyanide by air stripping were observed [15–17,39] and strongly complexed cyanide passed through the reaction basin. It is possible that in periods of upset condition, some cyanide air stripping might occur depending on the free cyanide concentration accumulated in the aeration basin. The amount of complexed cyanide accumulated in the sludge will depend on the SRT. In a Homestake mine wastewater treatment operation [46], the sludge age was relatively short. In treating coke plant wastewaters, operating sludge age tends to be long (40 to 100 days). At long SRTs (older microbial mass), less exopolymeric materials are produced as compared to younger sludges. This reduces the ability of the biomass to scavenge precipitated cyanide compounds. For any given study, it can be difficult to track the fate of cyanide species without careful consideration of many

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factors, including wastewater cyanide speciation, pH, aeration rate, development, and acclimation of the microbial mass, SRT, and sludge composition analyses.

23.2.2 ATTACHED GROWTH While suspended growth systems have achieved widespread usage in municipal and industrial wastewater treatment, biological processes involving microorganisms attached to surfaces are also popular and effective. Biofilm processes include trickling filters, rotating biological contactors, and fluidized beds. In each of these applications, microorganisms grow on a surface. Substrates diffuse into the biofilm and are transformed. Because the organisms are immobilized on surfaces, long cell retention times can be achieved without solids separation and recycling. 23.2.2.1 Trickling Filters The trickling filter is one of the earliest biological treatment processes developed, and is still widely used. The process entails the development of a microbial population that is attached to a support medium (e.g., slag, rock, or plastic). Modern trickling filters with plastic media packing have been built with packing depths varying from 4 to 12 m. The degradable materials in the wastewater are consumed by the biofilm. Hydraulic loading keeps the biofilm in a moist state. Co-current or countercurrent air circulation provides oxygen. Treated water and detached biomass are recovered at the bottom of the filter and are either sent to a clarifier to separate the cells, or discharged. Some treated water is recycled for hydraulic control. Figure 23.4 presents a schematic illustration of the trickling filter process. Tricking filters have plug-flow hydraulic regimes, leading to concerns regarding the toxic or inhibitory effects of constituents in the raw wastewater. For wastewaters with high cyanide or heavy metal concentrations, if insufficient dilution is available with the recycle stream, there may be toxicity or inhibition effects in the initial segment of the trickling filter. Pettet and Mills [4] first demonstrated cyanide acclimation in a trickling filter. Dicotr et al. [43] report on the evaluation of different media as supports for fixed-bed reactors designed to treat cyanide-containing wastewaters. Activated carbon, pozzolana, and a mixture of pumice stone and zeolite were all successful as substrates for cell attachment and growth. Cyanide was removed through a combination of biological reactions of abiotic precipitation reactions and adsorption processes. Dursun and Aksu [44] report on removal of dissolved ferrocyanide complex in a packedbed column reactor using Ca-alginate immobilized cells. Since dissolved iron–cyanide species are essentially not biodegradable (see Chapter 6), the removal process in this case, likely, was adsorption. 23.2.2.2 Rotating Biological Contactors A rotating biological contactor (RBC) system is a biofilm process that consists of a series of closely spaced circular plastic discs on a rotating shaft (see Figure 23.5). The discs/shaft assembly is partially submerged in a vessel containing wastewater and microbial material sloughed from the discs. Biomass accumulates on the discs, and as these discs rotate through the containing vessel, a film of water adheres to the biomass. This water film contains contaminants present in the wastewater, and the microbes that constitute the biofilm transform the contaminants. Although this biofilm can grow to significant thickness, actual microbial activity is limited to the outermost surface of the film. Microbes in the inner regions of the film are essentially deprived of substrate, nutrients, and oxygen by the activity in the surface layer. Oxygen is supplied by transfer from the air to the water film on the rotating discs. Patil and Paknikar [45] report on the treatment of silver cyanide in RBCs operating in continuous mode. The system achieved 99.5% removal of 0.1 mmol/L silver cyanide with molasses as a supplied

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Dome (optional) Protective surface grating

Hydraulic or motorized distributor

Cross flow media

Vertical flow media

Vent windows Blower (optional)

Media supports

Influent

Effluent Recycle

FIGURE 23.4 Trickling filter system components. Courtesy of Brentwood Industries.

carbon source. RBCs are also the reactor configuration at the Homestake Mine in Lead, SD site, where the first successful commercial biological cyanide treatment process for gold mining wastewaters was developed (see details later) [46].

23.2.2.3 Fluidized Bed Reactors Fluidized column reactors were developed to provide high-rate biological treatment. These reactors were designed to operate with very high microbial populations by growing the microorganisms on a support particle such as sand, activated carbon, or plastic bead. The microbial particles are fluidized by the upward flow of wastewater, air, and recycle water as shown in Figure 23.6. Oxygen is provided by the diffusion of either pure oxygen or air into the bottom of the column. Petrozzi and Dunn [47] report on successful free cyanide degradation in aerobic fluidized bed reactors. No growth with cyanide as sole carbon and nitrogen source was reported. Degradation with free cyanide as sole nitrogen source and an exogenous carbon source was complete, with the rate dependent upon oxygen transfer due to high chemical oxygen demand (COD) loading.

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Rotating bio-disks

Raw waste

Treated effluent Clarifier

Solids disposal

FIGURE 23.5 Schematic flow diagram of Rotating Biological Contractor process.

Effluent Recycle

Excess biological sludge

Influent

Sand biomass separation Sand

Oxygen or air Biofilm Media Bioparticle

FIGURE 23.6 Schematic of a fluidized bed biological treatment process.

23.2.3 NOVEL PROCESSES Chakraborty and Veeramani [48] report on a sequential anaerobic–anoxic–aerobic process to treat a synthetic wastewater that contains phenol, COD, cyanide, thiocyanate, and ammonia. Ye et al. [49] report that a constructed wetland removed selenocyanate (79 and 54% mass removal for selenium and cyanide, respectively). Patil and Paknikar [50] report on a combination biosorption/ biodegradation

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process with Cladosporium cladosporioides. The goal of the process was recovery and reuse of the metal cyanides. The process water was treatable by biodegradation with 99.9% efficiency.

23.2.4 ANAEROBIC PROCESSES Howe [51] first reported anaerobic removal of free cyanide. Fedorak and Hrudley [52] as well reported anaerobic treatment of free cyanide. Siller and Winter [53] report that free cyanide was removed in an anaerobic reactor with production of ammonia and formate. Biopass is an anaerobic system constructed in existing lined process solution ponds for the closure of heap leach pads [7]. The Biopass system is discussed in detail in Chapter 27. Treatment of free cyanide in batch anaerobic reactors was reported by Gijzen et al. [54]. They observed acclimation of methanogens to free cyanide influent levels of 125 mg/l. Removal of free cyanide was between 91 and 93% at loading rates of 250 mg/l-day. Cyanide inhibition was more pronounced for acetoclastic than for hydrogenotrophic methanogens. Annachhatre and Amornkaew [55] report treatment of a cyanide-containing wastewater in an upflow anaerobic sludge blanket (UASB) process. Free cyanide removal of 93 to 98% was reported for loadings of 0.38 kg/m3 -day (10 mg/l influent concentration).

23.2.5 PROCESS LIMITATIONS Many wastewaters that contain cyanide also contain other constituents that can interfere with biological processes. For example, Kumar et al. [56] report that oil contamination and poor sludge settleability are often observed in activated sludge plants that treat coke plant wastewater. As with any wastewater, considering biological treatment requires careful consideration of the composition of the waste and provisions for pretreatment (e.g., oil separation), if necessary, as illustrated by Kumar et al. [56]. For a wastewater that is strictly inorganic in character (e.g., from electroplating), biological treatment would be a poor treatment technology selection. Coke plant wastewater, which contains high concentrations of organics (e.g., 200 to 1300 mg/l phenols and 2,000 to 10,000 mg/l COD), is a good candidate for biological treatment. While one might expect gold mining tailings to be low in organic components, tramp materials (e.g., diesel fuel spills, degreasers, foaming lubricants, dispersants, biocides, and various chemical compounds) find their way into the tailings impoundment [24]. These materials could contribute to the organic content of the tailings wastewater. Experience at the Homestake Mine in Lead, South Dakota (see below) indicates that the combined tailings and mine water contain sufficient organic matter to allow for effective biological treatment.

23.3 FULL-SCALE APPLICATIONS OF BIOLOGICAL TREATMENT FOR CYANIDE IN GOLD MINING Biological treatment can be applied to gold mining wastes as deliberate or active treatment (i.e., biological reactors and associated equipment) or as passive incidental treatment, for example, natural degradation in the tailings ponds. Biological treatment of free and WAD cyanide in gold mining wastewaters has proven to be effective in meeting regulatory discharge limits and is more economical than physical-chemical treatment alternatives. In the mining industry, deliberate biological treatment of free and WAD cyanide has been effectively demonstrated at the Homestake Mine in Lead, South Dakota [24]. Homestake also has a biological treatment system at its Nickel Plate Mine in Canada [57]. The Homestake biological treatment facility in Lead, South Dakota, built in 1987, was designed to process 21,000 m3 /day (5.5 mgd) of mine water and tailings pond decant water. Table 23.1 presents the composition of the wastewater treated at this plant, and Figure 23.7 presents a schematic flow diagram of the treatment facility. There are two biological treatment stages: a cyanide treatment stage consisting of 24 RBC units with a total of 9,290 m2 (100,000 ft2 ) of surface area and 1 h hydraulic

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TABLE 23.1 Performance of the Homestake, Lead, SD, Biological Treatment Plant Feed wastewatera Parameter Thiocyanate Total cyanide WAD cyanide Total copper TSS Ammonia-N pH

Effluenta

Range

Averageb

Range

35–100 0.65–11.5 0.24–7.05 0.17–1.15 — 0.84–10.65 7.5–9.0

61.5 ± 6.1 3.67 ± 1.81 2.30 ± 1.88 0.56 ± 0.88 — 5.60 ± 1.20 —

— 0.03–1.12 0.02–0.41 0.01–0.13 1.0–18 0.5–0.90 6.25–8.35

Averageb 0.5 0.33 ± 0.15 0.05 ± 0.08 0.04 ± 0.02 3.0 ± 1.5 0.5 —

a All values are mg/l except pH in S.U. b Average includes ± standard deviation.

Source: Permission granted by Society for Mining Engineers: Whitlock, J., 1984, Proc. Society of Mining Engineers Annual Meeting, 1987.

retention time; and; a second, nitrification stage consisting of 24 RBC units with a total of 13,935 m2 (150,000 ft2 ) of surface area and 1.5 h of hydraulic retention time. Start up of the full-scale plant was initiated in 1984. Pseudomonas paucimobilis was the dominant organism. Nutrients were added, including inorganic carbon (soda ash) and phosphorous (H3 PO4 ). Removal rates were reported as 91 to 99.5% for total cyanide and 98 to 100% for WAD cyanide. Significant cost savings in capital and operating expenses compared with H2 O2 treatment were reported. In 1987, operating costs were 1.3% of total milling and mining costs at the site. The operating data at Homestake for 1988 show that the influent wastewater to treatment contained 62.0 mg/l SCN, 4.1 mg/l total CN, and 2.3 mg/l WAD CN. The treated effluent contained 0.37 mg/l total cyanide, 0.036 mg/l Method C cyanide, and 0.033 mg/l copper. Several other installations also operate full-scale biological treatment systems. The Green Spring Gold Operation in Ely, Nevada [58] utilizes attached growth on activated carbon columns. Peak Gold Mine in Western NSW, Australia, reported cyanide degradation in a process water basin equipped with surface aerators and nutrient dosing equipment [59]. The Nickel Plate Mine in British Columbia, Canada, uses a combined aerobic and anaerobic suspended growth biological treatment process to remove thiocyanate, ammonia, and nitrate [60]. While biological treatment has been implemented at a number of specific installations, it has not achieved widespread use for cyanide treatment in the gold mining industry.

23.4 BIOLOGICAL TREATMENT OF SOLID WASTES CONTAINING CYANIDE In addition to treatment of cyanide-bearing wastewaters, cyanide can be biodegraded in soils and groundwater. Kjeldsen [61] presents a comprehensive review of cyanide in soil and groundwater. While studies of the biodegradation of cyanide are extensive (as described in Chapter 6), very few studies have included soils or solids in the test matrices. Aronstein et al. [62] report on the role of microbiological processes on the removal of hexacyanoferrate from aqueous and soil-containing systems, including soils from manufactured gas plant sites. Amendment with several specific consortia of cyanide-degrading organisms, appeared to improve the extent of removal. Comparison of soil slurries with soil-free aqueous systems suggested that cyanide partitioning onto the solid phase reduces its availability for biodegradation.

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Tailings Decant Water

Flow Equalization System

Soda Ash

Rapid Mix Tank

Phosphoric Acid

Cyanide Treatment Stage Air

24 RBC Units 9,290 m 2 of Surface Area

Nitrification Stage Air

Polymer

24 RBC Units 2 13,935 m of Surface Area

Effluent Rapid Mix Tank

Effluent Rapid Mix Tank

Ferric Chloride

Emergency Clarifier

Multi-Media Pressure Sand Filtration

Discharge to Whitewood Creek

FIGURE 23.7 Process Flow Diagram for the Homestake Biological Wastewater Treatment Facility. (Source: Mudder, T., Fox, F., Whitlock, J., Fero, T., Smith, G., Waterland, R. and Vietl, J., in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd., London, 1998. With permission.)

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Cyanide-resistant bacteria are present in mine tailings, and bioremoval of radiolabeled spiked KCN to CO2 was observed in tailings [63]. Fresh mine tailings showed higher activity toward cyanide present in the tailings, while both fresh and old tailings transformed spikes of free cyanide [63]. Initial removal of cyanide from tailings is generally high, while over time the cyanide in old tailings appears to be more recalcitrant. The selective removal of free cyanide and persistence of metal–cyanide complexes which are not amendable to biotransformation, are implicated in this observation [64–66].

23.5 SUMMARY AND CONCLUSIONS • Biological treatment offers a cost-effective and reliable alternative for free and WAD cyanide treatment in a wide variety of wastewaters. • Microorganisms exist that are capable of degradation of free and WAD cyanide. Cyanide degrading organisms often exist together with nitrifying bacteria. • Cyanide can serve as a source of carbon and nitrogen for aerobic organisms. Nitrogen released in the aerobic biotransformation of cyanide can serve as an electron donor for nitrifying organisms. • Cyanide speciation controls biodegradability. System conditions including pH and presence of metals and complexing agents affect availability of cyanide to organisms. • Strong metal–cyanide complexes such as ferro- and ferricyanide species in water are not susceptible to biodegradation. • Inhibitory materials are a concern for the relatively slow-growing organisms involved in cyanide biodegradation. In a mixed culture of cyanide-degraders and nitrifiers, the sensitivity of nitrifiers is generally the most significant concern. Inhibitory levels of ammonia can build up if nitrifiers are inactive due to upset. • Contact time for biological processes is maintained by selection of HRT and MCRT. MCRT in commercial applications for cyanide biotreatment is generally quite long. • Suspended (batch, SBR, activated sludge) and attached (trickling filters, fluidized bed reactors, RBCs) growth systems have been studied for cyanide biotreatment systems. Most systems have been successful in the laboratory, while few have been scaled up for commercial applications. • In gold mining, biotreatment of free and WAD cyanide has been implemented at full scale at some locations but has not been widely adopted. A successful full-scale treatment system for mine wastes is at the Homestake Mine in Lead, South Dakota. • In coke plant wastewater treatment systems, biological treatment is a widely used process for removal of ammonia, cyanide, and thiocyanate that remain after ammonia stripping. • A few studies showing biotransformation of free and WAD cyanide in solid wastes have been performed.

REFERENCES 1. Velz, C.J., A basic law for the performance of biological filters, Sewage Works J., 20, 607, 1948. 2. Arden, E. and Lockett, W.T., Experiments on the oxidation of sewage without the aid of filters, J. Soc. Chem. Indust., 33, 523, 1914. 3. Rittmann, B.E. and McCarty, P.L., Environmental Biotechnology: Principles and Applications, McGraw-Hill, New York, 2001. 4. Pettet, A.E.J. and Mills, E.V., Biological treatment of cyanides, with and without sewage, J. Appl. Chem., 4, 434, 1954. 5. Neufeld, R., Greenfield, J., and Rieder, B., Temperature, cyanide and phenolic nitrification inhibition, Water Res., 20, 633, 1986.

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6. Akcil, A. and Mudder, T., Microbial destruction of cyanide wastes in gold mining: process review, Biotechnol. Lett., 25, 445, 2003. 7. Akcil, A., Karahan, A.G., Ciftci, H., and Sagdic, O., Biological treatment of cyanide by natural isolated bacteria (Pseudomonas sp.), Miner. Eng., 16, 643, 2003. 8. Richardson, K.R., European Union Patent EP 233719, August 26, 1987. 9. Richardson, K.R. and Clark, P.M., European Union Patent EP 234760, 1987. 10. Ingvorsen, K., European Union Patent 349348A1, March, 1990. 11. Karavaiko, G.I., Kondrat’eva, T.F., Savari, E.E., Grigor’eva, N.V., and Avakyan, Z.A., Microbial degradation of cyanide and thiocyanate, Microbiology, 69, 167, 2000. 12. Dursun, A.Y., Calik, A., and Aksu, Z., Degradation of ferrous (II) cyanide complex ions by Pseudomonas fluorescens, Process Biochem., 34, 901, 1999. 13. Mudder, T.I., Wastewater treatment using acclimatized Pseudomonas paucimobilis, Gold and Silver Recover Innovations, Phase III, 9, 5390, 1987. 14. Knowles, C.J., Microorganisms and cyanide, Bacteriol. Rev., 40, 652, 1976. 15. Wong-Chong, G.M. and Caruso, S., Biological treatment of by-product coke plant wastewater for the control of BAT parameters, in Proc. Symp. Iron and Steel Pollution Abatement Technology for 1981, U.S. Environmental Protection Agency, Washington, DC, 1982. 16. Wong-Chong, G.M. and Caruso, S.C., Advanced biological oxidation of coke plant wastewaters for the removal of nitrogen compounds, Carnegie Mellon Research Institute report to the American Iron and Steel Institute, Pittsburgh, PA, 1977. 17. Wong-Chong, G.M., Biological degradation of cyanide in complex industrial wastewaters. Proc. Internat. Symp. Biohydrometallurgy, BIOMINET, CANMET Mining and Mineral Sciences Laboratories, Natural Resources Canada, Jackson Hole, WY, 289, 1989. 18. Anthonisen, A.C., Loehr, R.C., Prakasam, B.S., and Srinath, E.G., Inhibition of nitrification by ammonia and nitrous acid, J. Water Poll. Control Fed., 48, 835, 1976. 19. Sharma, B. and Ahlert, R.C., Nitrification and nitrogen removal, Water Res., 11, 897, 1977. 20. Luthy, R.G. and Jones, L.D., Biological oxidation of coke plant effluent, J. Environ. Eng. Div. ASCE, 106, 847, 1980. 21. Luthy, R.G., Treatment of coal coking and coal gasification wastewaters, J. Water Poll. Control Fed., 53, 325, 1981. 22. Thiem, L.T. and Alkhatib, E.A., In situ adaptation of activated sludge by shock loading to enhance treatment of high ammonia content petrochemical wastewater, J. Water Poll. Control Fed., 60, 1245, 1988. 23. WEF, Pretreatment of Industrial Wastes (MOP FD-3); WEF Manual of Practice Manuals and Reports on Engineering Practice, Water Environment Federation, 1994. 24. Whitlock, J.L. and Smith, G.R., Operation of Homestake’s cyanide biodegradation wastewater system based on multi-variable trend analysis, Proc. Int. Symp. Biohydrometallurgy, Jackson Hole, WY, 613, 1989. 25. Sorokin, D.Y., Tourova, T.P., Lysenko, A.M., and Kuenen, J.G., Microbial thiocyanate utilization under high alkaline conditions, Appl. Environ. Microbiol., 67, 528, 2001. 26. Wong-Chong, G.M., Caruso. S.C., and Patarlis, T.G., Treatment and control technology for coke plant wastewaters, in Proc. AICHE 84th National Meeting, Atlanta, GA, 1978. 27. Knowles, C.J. and Bunch, A.W., Microbial cyanide metabolism, Adv. Microbial. Physiol., 27, 74, 1986. 28. VanBriesen, J.M., Rittmann, B.E., Girvin, D., and Bolton, H. Jr., The rate-controlling substrate form for the biodegradation of Nitrilotriacetate by Chelatobacter heintzii, Environ. Sci. Technol., 34, 3346, 2000. 29. Finnegan, I., Toerien, S., Abbot, L., Smith, F., and Raubenheimer, H.G., Identification and characterization of an Acinetobacter sp. capable of assimilation of cyano-metal complexes, free cyanide ions and organic nitriles, Appl. Microbiol. Biotechnol., 36, 142, 1991. 30. Adjei, M.D. and Ohata, Y., Factors affecting the biodegradation of cyanide by Burkholderia cepacia Strain C-3, J. Biosci. Bioeng., 89, 274, 2000. 31. Dumestre, A., Chone, T.P.J.M., Gerard, M., and Berthelin, J., Cyanide degradation under alkaline conditions by a strain of Fusarium solani isolated from contaminated soils, Appl. Environ. Microbiol., 63, 2729, 1997.

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32. Shivaraman, N. and Parhad, N.M., Biodegradation of cyanide in a continuously fed aerobic system, J. Environ. Biol., 5, 273, 1984. 33. Boucabeille, C., Bories, A., Ollivier, P., and Michel, G., Microbial degradation of metal complexed cyanides and thiocyanate from mining wastewaters, Environ. Pollut., 84, 56, 1994. 34. White, D.M., Pilon, T.A., and Woolard, C., Biological treatment of cyanide containing wastewater, Water Res., 1, 2105, 2000. 35. Richards, D.J. and Shieh, W.K., Anoxic oxic activated sludge treatment of cyanides and phenols, Biotechnol. Bioeng., 33, 32, 1989. 36. Mihaylov, B.V. and Hendrix, J.L., Biological decomposition of cyanide in sewage sludge, Miner. Eng., 7, 61, 1994. 37. Raef, S.F., Characklis, W.G., Kessick, M.A., and Ward, C.H.,. Fate of cyanide and related compounds in aerobic microbial systems. I. Chemical reaction with substrate and physical removal, Water Res., 11, 477, 1977. 38. Raef, S.F., Characklis, W.G., Kessick, M.A., and Ward, C.H., Fate of cyanide and related compounds in aerobic microbial systems. II. Microbial degradation, Water Res., 11, 485, 1977. 39. Harden, D., Jones, D.D., and Gauthier, J.J., Adaptation of an industrial activated sludge process to the removal of cyanide, in Proceedings of the 38th Purdue Industrial Waste Conference, Purdue University, West Lafayette, IN, 289, 1983. 40. Pandey, R.A., Kumaran, P., Shivaraman, R., Parahad, N.M., and Kaul, S.N., Cyanide and thiocyanate removal from LTC Wastewater, Asian Environ., 9, 10, 1987. 41. Neufeld, R.D., Mattson, L., and Lubon, P., Thiocyanate biooxidation kinetics, J. Environ. Eng. Div. — ASCE, 107, 1035, 1981. 42. Desai, J.D. and Ramakrishna, C., Microbial degradation of cyanides and its commercial applications, J. Sci. Indust. Res., 57, 441, 1998. 43. Dictor, M.-C., Battaghlia-Brunet, F., Morin, D., Bories, A., and Clarens, M., Biological treatment of goal ore cyanidation wastewater in fixed bed reactors, Environ. Pollut., 97, 287, 1997. 44. Dursun, A.Y. and Aksu, Z., Biodegradation kinetics of ferrous(II) cyanide complex ions by immobilized Pseudomonas fluorescens in a packed bed column reactor, Proc. Biochem., 35, 615 , 2000. 45. Patil, Y.B. and Paknikar, K.M., Biodetoxification of silver–cyanide from electroplating industry wastewater, Lett. Appl. Microbiol., 30, 33, 2000. 46. Whitlock, J. and Mudder, T., The Homestake wastewater treatment process: Biological removal of toxic parameters from cyanidation wastewaters and bioassay effluent evaluation, in Cyanide Monograph, Mudder, T. I., Ed., Mining Journal Books Ltd., London, U.K., 1998. 47. Petrozzi, S. and Dunn, I.J., Biological Cyanide degradation in aerobic fluidized bed reactors: treatment of almond seed wastewater, Bioproc. Eng., 11, 29, 1994. 48. Chakraborty, S. and Veeramani, H., Anaerobic–Anoxic–Aerobic Sequential Degradation of Synthetic Wastewaters, Appl. Biochem. Biotechnol., 102–103, 443, 2002. 49. Ye, Z.H., Lin, Z.-Q., Whiting, S.N., de Souza, M.P., and Terry, N., Possible use of constructed wetland to remove selenocyanate, arsenic, and boron from electric utility wastewater, Chemosphere, 52, 1571, 2003. 50. Patil, Y.B. and Paknikar, K.M., Removal and Recovery of metal cyanides using a combination of biosorption and biodegradation processes, Biotechnol. Lett., 21, 913, 1999. 51. Howe, R.H.L., Biodestruction of cyanide wastewater — advantages and disadvantages, Air Water Pollut., 9, 463, 1965. 52. Fedorak, P.M. and Hrudey, S.E., Cyanide transformation in anaerobic phenol-degrading methanogenic cultures, Water Sci. Technol., 21, 67, 1989. 53. Siller, H. and Winter, J., Degradation of cyanide in agroindustrial or industrial wastewater in an acidification reactor or in a single-step methane reactor by bacteria enriched from soil and peels of cassava, Appl. Microbiol. Biotechnol., 50, 384, 1998. 54. Gijzen, H.J., Bernal, E., and Ferrer, H., Cyanide toxicity and cyanide degradation in anaerobic wastewater treatment, Water Res., 34, 2447, 2000. 55. Annachhatre, A.P. and Amornkaew, A., Upflow anaerobic sludge blanket treatment of starch wastewater containing cyanide, Water Environ. Res., 73, 622, 2001.

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56. Kumar, M.S., Vaidyan, A.N., Shivaraman, N., and Bal, A.S., Biotreatment of oil-bearing coke-oven wastewater in fixed film reactor: a viable alternative to activated sludge process, Environ. Eng. Sci., 17, 221, 2000. 57. Whitlock, J.L. and Whitlock, C.W., Recent advances in technologies for biological treatment of thiocyanate, cyanide, heavy metals and nitrates, in Cyanide: Social, Industrial, and Economic Aspects, Young, C.A., Twidwell, L.G., Anderson, C.G., Eds., The Minerals, Metals and Materials Society, 195, 2001. 58. Dubey, S.K. and Holmes, D.S., Biological cyanide destruction mediated by microorganisms, World J. Microbiol. Biotechnol., 11, 257, 1995. 59. Bernoth, L., Firth, I., McAllister, P., and Rhodes, S., Biotechnologies for remediation and pollution control in the mining industry, Min. Mettalurg. Proc., 17, 105, 2000. 60. Given, B., Dixon, B., Douglas, G., Mihoc, R., and Mudder, T., Combined aerobic and anaerobic biological treatment of tailings, solution at the nickel plate mine, in Cyanide Monograph, Mudder, T.I., Ed., Mining Journal Books, Ltd., London, U.K., 1998. 61. Kjeldsen, P., Behaviour of cyanides in soil and groundwater: a review, Water, Air, Soil Pollut., 115, 279, 1999. 62. Aronstein, B.N., Maka, A., and Srivastava, V.J., Chemical and biological removal of cyanides from aqueous and soil-containing systems, Appl. Microbiol. Biotechnol., 41, 700, 1994. 63. Oudjehani, K., Zagury, G.J., and Deschenes, L., Natural attenuation potential of cyanide via microbial activity in mine tailings, Appl. Microbiol. Tiotechnol., 58, 409, 2001. 64. Laha, S. and Luthy, R.G., Investigation of microbial degradation of fixed cyanide, Report to Aluminum Association and Gas Research Institute, Carnegie Mellon University, Pittsburgh, PA, 1991. 65. Zagury, G.J., Oudjehani, K., and Deschenes, L., Characterization and availability of cyanide in solid mine tailings from gold extraction plants, Sci. Total Environ., 320, 211, 2004. 66. Whitlock, J., Performance of the Homestake mining company biological cyanide degradation wastewater treatment plant August 1984. Proc. Society of Mining Engineers Annual Meeting, 1987.

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24 Cyanide Phytoremediation Stephen D. Ebbs, Joseph T. Bushey, Brice S. Bond, Rajat S. Ghosh, and David A. Dzombak CONTENTS 24.1

Phytoremediation: An Emerging Biotechnology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.1.1 Phytoremediation Biotechnologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.1.2 Cyanide Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.2 Cyanide Contamination in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.3 Feasibility of Cyanide Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.3.1 Cyanide Metabolism and Toxicity in Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.3.2 Cyanide Phytoremediation Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.3.2.1 Phytoremediation of Free Cyanide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.3.2.2 Phytoremediation of Complexed Cyanide Compounds . . . . . . . . . . . . . 24.4 Modeling Cyanide Uptake and Fate in Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.4.1 Model Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.4.2 Model Calibration and Simulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.5 Cyanide Phytoremediation: Limitations and Challenges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.5.1 Selection of Plant Species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.5.1.1 General Plant Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.5.1.2 Plant Characteristics Specific for Cyanide Phytoremediation . . . . . . 24.5.2 Use of Multiple Species. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.6 Possible Scenarios for Cyanide Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.6.1 Engineering Implications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.6.2 Field Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.7 Regulatory Concerns . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.7.1 Cyanide Bioaccumulation in Plant Tissue . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.7.2 Cyanide Volatilization from Plant Tissues. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24.8 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

480 480 481 481 482 482 483 483 485 486 488 489 490 490 490 491 492 492 493 494 495 495 496 497 498

Since the advent of the Industrial Age, anthropogenic activities have introduced a wide variety of contaminants into the environment. While heavy metals, PCBs, PAHs, dioxins, and pesticides have received significant attention, other contaminants, including cyanide, are emerging as concerns. In the United States, cyanide has been among the top 30 of the 275 compounds on the CERCLA Priority List of Hazardous Substances since 1995 and also appears on the list of compounds frequently found at Superfund sites. Cyanide is also included in water, wastewater, and hazardous waste regulations as discussed in Chapter 18. The need for cost-effective methods has grown with the need for treatment and remediation of cyanide contamination. Phytoremediation is the use of vascular plants, algae, and fungi to metabolize or sequester contaminants, or to induce contaminant breakdown by micro-organisms in soil [1]. Green plant 479

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transformation and sequestration of contaminants has received considerable attention in recent years as an alternative or an augmentation to physically based treatment approaches. The use of plants for remediation is projected to be less expensive than contemporary methods, but the economics of phytoremediation depends on a number of project-specific factors, including the availability and cost of land. While phytoremediation is a promising technology, progress toward its implementation for cyanide is limited by the lack of fundamental knowledge of the physicochemical mechanisms that influence cyanide fate processes in the environment, biological mechanisms responsible for the transport and metabolism of cyanide species by plants, and cyanide uptake rates and factors that influence these rates (e.g., types and density of plants). Reviewed here are recent laboratory and field studies of phytoremediation of cyanide species, including free cyanide and metal-complexed cyanide compounds (e.g., ferrocyanide). Also discussed herein are the environmental factors that influence cyanide bioavailability, speciation, and, ultimately, the efficacy of phytoremediation in the field. A plant-scale phytoremediation process model for cyanide is outlined, and example applications involving a constructed wetland scenario for the remediation of cyanide-contaminated water are presented. Additional scenarios where cyanide phytoremediation could be utilized are also discussed. Finally, factors related to the regulatory acceptance of phytoremediation, such as cyanide bioaccumulation and volatilization, are discussed.

24.1 PHYTOREMEDIATION: AN EMERGING BIOTECHNOLOGY 24.1.1 PHYTOREMEDIATION BIOTECHNOLOGIES Phytoremediation is a family of emerging biotechnologies that utilize plants for the remediation of environmental contamination. It includes five primary technologies: phytoextraction, rhizofiltration, phytostabilization, phytovolatilization, and phytodegradation [2]. Phytoextraction is the use of plants to remove metals, metalloids, or other contaminants from soil and concentrate those contaminants in above-ground plant tissues. The contaminants are removed from the site by harvesting the aerial tissues. The primary advantages of phytoextraction are the reduction in the volume of waste to be stored (harvested plant material vs. a larger volume of excavated soil) and the potential recovery of economically important metals (e.g., zinc) from the harvested plant material (“biomining”). Limitations include the achievable plant concentrations in the aerial tissues, contaminant bioavailability, and the biomass production of the appropriate plants. Rhizofiltration is the use of plant roots to remove contaminants from polluted waters. This can be performed on small bodies of water, such as ponds, under ambient conditions using floats that support plants with prolific roots. Sunflower (Helianthus annus L.) and Indian mustard (Brassica juncea L.) have been the species most often used in this context [3,4]. Alternatively, rhizofiltration can be the treatment part of a pump-and-treat approach in which contaminated surface or groundwater is pumped through a system of plant roots. As the cell walls of plant roots are negatively charged and tend to have a large sorption capacity, rhizofiltration can be highly effective in the removal of cationic contaminants. While not specifically considered a rhizofiltration approach, constructed wetlands operate under much the same principle in that the contamination is removed from an aqueous medium by plants, rather than from soil. Constructed wetlands are more complex engineered structures that have been used to remediate a wide variety of contaminants including heavy metals [5], metalloids (e.g., selenium [6,7]), and a range of organic compounds [8–11]. Phytostabilization is the use of plants to reduce the solubility of contaminants in soils, primarily through modification of the physicochemical conditions in the rhizosphere (e.g., pH or pE), thereby reducing contaminant solubility, mobility, or toxicity. This has been investigated for chromium contamination [12], and is also being applied to uranium contamination. Another aspect of phytostabilization involves the use of metal-tolerant grasses that preferentially sequester metals in

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their roots. Since plants used for phytostabilization are selected because they have a low shoot accumulation of the contaminants, the above-ground biomass can be managed using traditional techniques and is not considered as a hazardous waste [13]. The root material is typically left in place. Phytovolatilization uses plants that remove contaminants from terrestrial or aqueous systems and facilitate their conversion to volatile forms for release to the atmosphere. In some cases, the volatile form of the contaminant is less toxic than the parent form. Phytovolatilization has been proposed as a means of remediating selenium contamination [14,15] as well as some volatile organic compounds such as trichloroethylene [16,17]. Biotechnology has as well been used to develop plants capable of volatilizing mercury [18], but the regulatory and public acceptance of this approach is still a matter of debate. Finally, phytodegradation involves the use of plants to metabolize mutable contaminants, including hydrocarbons, ammunition wastes, chlorinated solvents, and herbicides. Phytodegradation may be a precursor to phytovolatilization, with the metabolic products more volatile than the parent compound, or may act in parallel. The extent of degradation varies by contaminant, with some studies showing incomplete degradation of the contaminant [19]. The prospects for complete contaminant degradation and assimilation by plants have not been fully explored but are certainly germane to the implementation of phytodegradation in the field, as well as to regulatory acceptance of this approach.

24.1.2 CYANIDE PHYTOREMEDIATION Cyanide contamination is a new target for phytoremediation. Cyanide phytoremediation shows promise as there is already extensive evidence of cyanide metabolism in plants (see Chapter 6). Based upon results obtained from preliminary studies, cyanide phytoremediation could encompass all of the plant-related biotechnologies described above, as cyanide contamination is present in both terrestrial and aqueous systems. The most exciting prospect for cyanide phytoremediation is the possibility of cyanide phytodegradation. Since many organisms, including plants, assimilate cyanide completely into primary metabolism, cyanide phytoremediation could be one of the few instances in which phytoremediation results not only in contaminant removal, but also in contaminant destruction. If assimilated and used constructively, cyanide contaminants could ultimately be beneficial for plant growth and development. However, some evidence of cyanide volatilization by plants exists (see Chapter 6), raising regulatory concerns if it is shown to be a primary route of cyanide disposition by plants. The regulatory implications of cyanide phytovolatilization are discussed in more detail in Section 24.7.2.

24.2 CYANIDE CONTAMINATION IN THE ENVIRONMENT As discussed in Chapters 2 and 5, cyanide exists in the environment in a variety of forms, including various complexes. Free cyanide released into the environment complexes readily with soil cations such as Cd2+ , Co2+ , Cu2+ , Fe2+/3+ , and Ni2+ to form metal–cyanide complexes [20]. The degree of dissociation of these metal–cyanide complexes into free cyanide (CN− or HCN) can vary greatly. Metal–cyanide complexes such as cobalt and iron require extreme acidity for dissociation and are termed “strong acid dissociable complexes,” whereas complexes such as nickel, copper, silver, zinc, and cadmium require mild acidity to dissociate and are termed “weak acid dissociable complexes.” Because of the weak bonding between the cyanide ion and the corresponding metal, the “weak acid dissociable complexes” ultimately can yield more free cyanide in the environment. However, the most common forms of metal–cyanide complexes found in the environment are the iron–cyanide complexes such as ferrocyanide [Fe(CN)6 ]−4 and ferricyanide [Fe(CN)6 ]−3 complexes, because of the abundance of iron in soil and the reactivity of the CN− ion. The strongly bonded nature of

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these complexes makes them less susceptible to biodegradation (Chapter 6) and they usually require energy-intensive processes for treatment. Iron–cyanide complexes, collectively referred to as hexacyanoferrates, can precipitate under excess iron and neutral to acidic pH conditions to form Prussian blue [Fe4 (Fe(CN)6 )3 (s)] or Turnbull’s blue [Fe3 (Fe(CN)6 )2 (s)]. These blue solids produce the characteristic blue color observed in spent oxide box wastes at former manufactured gas plant (MGP) sites, and in spent potliner wastes from aluminum smelters. In the past, the blue pigments were used extensively in dyes, inks, pharmaceuticals, and cosmetics. Iron cyanide compounds are also used as anticaking agents in road salt, primarily in the form of sodium ferrocyanide or yellow prussiate of soda (YPS) [21]. Details regarding the speciation of cyanide and the species-specific physicochemical properties are discussed in Chapters 2 and 5.

24.3 FEASIBILITY OF CYANIDE PHYTOREMEDIATION 24.3.1 CYANIDE METABOLISM AND TOXICITY IN PLANTS Cyanide phytoremediation will be technically feasible only if plants can remove and subsequently metabolize the cyanide compounds without succumbing to cyanide toxicity. Ample evidence exists of cyanide metabolism in plants (Chapters 3 and 6) through the activity of enzymes such as β-cyanoalanine synthase and sulfur transferases (rhodanese-like enzymes). Furthermore, the plant mitochondrial alternative oxidase, minimizes the inhibition of electron transport, a characteristic of cyanide intoxication. These pathways act in concert to provide cyanide homeostasis within plant tissues. Plants that are cyanogenic, for instance, produce cyanogenic glycosides as a defensive compound and storage form for nitrogen (Chapter 3). Yet, free cyanide concentrations are typically maintained at low levels in such plants. The sensitivity of plants to cyanide intoxication, and hence their efficacy for cyanide phytoremediation, will be dependent upon the balance between the activity of these metabolic pathways, and the concentration and duration of cyanide exposure. For example, the mode of toxic action for the auxin herbicide quinclorac involves the stimulation of excess cyanide production in sensitive plants, resulting in an increase in cyanide concentration in the tissues [22]. Tolerant plants do not show this increase. Few studies have determined the concentration dependence of cyanide toxicity in plants. External cyanide concentrations of 1 µM (26 µg l−1 as free cyanide) are considered nontoxic for Mouseear cress (Arabidopsis thaliana L.) [23]. Bush bean (Phaseolus vulgaris L.) showed pronounced wilting after a 3-day exposure to 25 mg NaCN kg−1 soil (DW). A complete loss of whole plant turgor caused bush beans to collapse 1 day after the addition of 100 mg NaCN kg−1 soil (DW), with 100% mortality after 9 days [24]. Studies have shown willow (Salix spp.), a promising plant for cyanide phytoremediation, to be more resilient. Using transpiration as an assay for toxicity, the EC10 for a 72 h exposure to cyanide (CN− ) was determined to be 0.3 mg l−1 for basket willow (Salix viminalis L.) [25], with an EC50 value of 1.5 mg l−1 . In contrast, diamond willow (Salix eriocephala var. Michaux) showed no change in transpiration or growth as compared to controls following a 20-day exposure to 2 mg l−1 free cyanide [26]. Thiocyanate is generally regarded as less toxic to animals than cyanide. The same contention seemingly holds true for plants as there is little evidence of thiocyanate toxicity. Wheat (Triticum aestivum L.) treated with ammonium thiocyanate at a rate of ∼900 kg ha−1 showed no adverse effects and expressed a stimulation of growth after 69 days [27]. The fate of the thiocyanate in wheat was not described. Plants also use thiocyanate as a central component of the glucosinolate metabolism in the Brassicaceae family (e.g., cabbage, broccoli, turnip, and Indian mustard), occurring as both SCN− and thiocyanate glucosides. Metal cyanide compounds such as ferrocyanide can also be toxic to plants, although the mechanism by which this occurs is unclear. Several studies have shown that roots from plants exposed to hexacyanoferrates (principally ferrocyanide), can become coated with iron cyanide solids such

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as Prussian blue and Turnbull’s blue [28,29], as has been observed for diamond willow, tomato (Lycopersicum esculentum L.), Indian mustard (Brassica juncea L.), and various grasses. Similar precipitates can be observed on the roots of plants growing on soils contaminated with metal–cyanide complexes. Introduction of 8 mg l−1 ferrocyanide into hydroponic nutrient solution resulted in heavy precipitation of complexed cyanide solids on the roots of diamond willow [28]. The willow plants showed decreased water use, biomass, and leaf area, as compared to controls. Concentrations of ferrocyanide between 30 and 143 mg l−1 resulted in similar precipitation for grasses such as foxtail (Setaria sp.), barley (Hordeum vulgare L.), and wild cane (Sorghum bicolor L.), following a 14 to 28 days exposure [30]. Diamond willow grown under the same conditions showed similar precipitation but a smaller decrease in biomass [28]. It is unclear whether these hexacyanoferrates are themselves toxic, or whether the resulting adverse effects were indirectly based upon the blockage of water and solute movement across the root cell walls. Thermodynamics predicts the formation of cyanide solids under such concentrations at equilibrium conditions. Removal of soluble iron sources from the hydroponic media prevented precipitation of ferrocyanide [26,31]. There was no loss of biomass for barley, oat, or wild cane as compared to controls, allowing for exposures of up to 50 mg ferrocyanide L−1 , with no adverse effects [31]. Balsam poplars (Populus trichocarpa L.) showed reduced growth but survived ferric ferrocyanide (Prussian blue) concentrations of up to 2500 mg l−1 [24]. These studies suggest a tentative range of concentrations over which cyanide phytoremediation may be feasible, based upon innate resistance of plants to toxicity from various cyanide species. Further study will be necessary to determine an effective concentration range that can be phytoremediated without risk of phytotoxicity relative to the plant species selected. Plant tolerance and uptake of cyanide depends on cyanide speciation and interaction with the soil, which varies with the physicochemical conditions of the media, and will therefore influence phytotoxicity and remedial effectiveness. The presence of iron reduced cyanide solubility, altered solution speciation, and resulted in decreased plant biomass. Other factors, such as pH, redox potential, and the relative solution concentrations are additional variables that affect cyanide speciation.

24.3.2 CYANIDE PHYTOREMEDIATION STUDIES Only a handful of studies have investigated the technical feasibility of cyanide phytoremediation. Most have been conducted in hydroponic studies in greenhouses, but some have also generated field data [32]. All the results of these studies generally support the prospect of cyanide phytoremediation, in that they have provided valuable physiological data that will be needed to understand the technology and successfully implement phytoremediation in the field. 24.3.2.1 Phytoremediation of Free Cyanide The majority of cyanide phytoremediation studies have focused on removal and metabolism of cyanide by willow (Salix sp.) and poplar (Populus sp.). Experiments using free cyanide have shown this chemical species to be readily removed from hydroponic media by intact willow plants, willow cell suspensions, and even desiccated willow plants [24,26,33]. For example, 75 to 90-day-old diamond willows (Salix eriocephala var. Michaux), established from cuttings obtained from a plant growing in cyanide-contaminated soil, removed 95% of free cyanide (as CN) mass from a hydroponic solution initially containing 2 mg l−1 total cyanide as CN in 20 days [26]. Hydroponic studies with white willow (Salix alba L.) generated similar results, with reported concentration ratios (mg CN kg FW−1 plant/mg CN l−1 solution) of 2.0 to 3.0 for roots, 0.2 for stems, and 0.1 to 0.6 for leaves following a 24-h exposure [25]. Concentration ratios (CR) are often used in phytoremediation as a measure of efficiency, with values greater than one indicative of a plant species capable of accumulating the contaminant of interest to concentrations greater than that observed in the medium.

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The application of the CR concept to cyanide phytoremediation has an important limitation. In the aforementioned study with diamond willows, the cyanide concentration in all tissues was <1 mg kg FW−1 at day 20 [26]. If CR values were calculated from solution and tissue concentrations, the day 20 values would be 0.56 for roots, 0.02 for stems, and 0.12 for leaves, which are not impressive as indicators of the cyanide removal efficiency. However, in addition to chemical analyses for cyanide content, the study with diamond willow used cyanide labeled with the stable isotope 15 N, providing a means of tracing the movement of the cyanogenic nitrogen atom within the plant system, and also of correlating the presence of 15 N with the tissue cyanide content. While tissue cyanide concentrations were <1 mg kg FW−1 , 15 N concentrations ranged from 250 to 2500 mg kg FW−1 in roots, stems, and leaves, representing increases in the 15 N abundance of 5 to 10% for roots, 2 to 3% for stems, and <1% for leaves. Given the natural abundance of 15 N in plant tissues, ∼0.37%, the increased 15 N abundances reflect a significant uptake and transport of cyanogenic nitrogen by willow, without a corresponding increase in tissue cyanide concentration. Due to the assimilation of cyanide by plants, CR values cannot be applied in this context to evaluate the effectiveness of cyanide uptake or phytoremediation. Such ratios are less meaningful as they fail to account for the potential metabolism of cyanide by the plant. The difference in concentration ratios between the studies with white and diamond willow also implies that cyanide concentration in the tissue may be affected by time or plant species. In either case, the measurement of tissue cyanide concentration should not be excluded from phytoremediation studies of this type because of the potential regulatory implications. A 0.04-acre pilot-scale constructed wetland at Blue Pond in Alcoa, TN, consisting of two openwater tandem cells (75 and 87 m2 , respectively), has demonstrated removal of free cyanide from contaminated water under field conditions [32]. This constructed wetland contained both planted vegetation, including softstem bulrush (Scirpus validus L.) and cattails (Typha sp.), and volunteer vegetation, primarily coontail (Ceratophyllum demersum L.) and pondweeds (Potamageton spp.). A significant reduction (56%) in free cyanide was observed between the inflow (influent free cyanide concentration of 0.005 to 0.02 mg l−1 and an influent flow rate of 1 gpm) and outflow, with free cyanide concentrations in the outflow consistently less than the laboratory reporting limit of 0.005 mg free cyanide l−1 . The mechanism responsible for the removal of cyanide from the contaminated water was not clear. Volatilization of cyanide was suggested, but could not be confirmed with subsequent experiments. Sediment samples showed increased free and total cyanide concentrations relative to the inflow, but the analyses were complicated by interferences in the samples. However, calculations based upon in vitro experiments with micro-organisms isolated from the wetland suggested that accumulation of cyanide in sediments was not likely to be the primary means by which cyanide was removed from the inflow water. If volatilization and deposition were not responsible for cyanide decreases, removal by plants and other organisms present in the wetland represents the only remaining removal mechanism for cyanide. Another pilot-scale wetland facility, Duck Spring, located near Blue Pond at Alcoa, TN, was constructed to treat free cyanide and iron–cyanide complexes [34]. The Duck Spring pilot-scale wetland covers an area of 0.5 acre and is constructed to treat ∼10 gpm flow diverted from a limestone groundwater spring impacted with waste leachate from an aluminum smelting facility. Primary emergent plant species at this wetland facility include river bulrush, softstem bulrush, sweet flag, common arrowhead, and American bur-reed. In addition, there are submergent species, such as coontail. Figure 24.1 provides a schematic of the Duck Spring wetland facility. The total cyanide concentration in the wetland influent is ∼0.3 ppm with the cyanide speciation dominated by the presence of strong cyanide complexes, that is, ferrocyanide (0.27 ppm), while free cyanide constitutes ∼10% of the total cyanide (0.03 ppm). Results from the first 21 days of intense monitoring that took place in July 2004, seven months after the wetland was put into operation, indicated 97% removal of total cyanide and 100% removal of free cyanide in the wetland, operating at a hydraulic retention time (HRT) of ∼5.6 days. For perspective, the total cyanide mass loss rate during the 21-day monitoring event amounted to 0.073 kg day−1 .

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Springfield Creek Discharge to creek

Weir Duck Spring inlet

Collection pipe Flow

27
–5

Outlet 4”PVC discharge

Spring Orifice (200 gpm)

D2, 12 I.D.PVC

Flow

27
–5 AgriDrain out

AgriDrain in Flow Distributor Pipe, 4 PVC perforated

~10 gpm 264


27
–5

D1
12 I.D. PVC Area: 0.5 acre

FIGURE 24.1 Schematic of Duck Spring Pilot-Scale Wetland located near Alcoa, TN. (Source: GTI, Report No. 8643, Gas Technology Institute, Des Plaines, IL, 2004. With permission.)

24.3.2.2 Phytoremediation of Complexed Cyanide Compounds As discussed in Chapters 2 to 5, complexed cyanide compounds can be the predominant forms of cyanide present in many environmental matrices. Although free cyanide has low cost, biological treatment alternatives, metal–cyanide complexes usually require energy-intensive, high-cost treatment options. From the economic standpoint, phytoremediation, if proven successful, could be an attractive technology for treating metal complexed cyanide species when sufficient low-cost land is available for implementation. Furthermore, because of the dominance of complexed cyanide in many environmental matrices, it is likely that phytoremediation efforts will have to contend with this chemical species more frequently than free cyanide. The few studies that have examined plant uptake and transport of complexed cyanide compounds (principally ferrocyanide) suggest that phytoremediation can be applied to these compounds, but that the rate of removal is considerably slower than that observed for free cyanide. In one study, cuttings of diamond willow were exposed to 6 mg l−1 of 15 N-labeled iron cyanide (95% ferrocyanide) in hydroponic solution. After 3 weeks of treatment, isotopic analysis of the plants indicated that the cyanogenic nitrogen (15 N) had been translocated to leaves. Treated plants showed a 500-fold greater 15 N enrichment than control plants, with cyanogenic nitrogen accounting for ∼0.5% of total leaf N [28]. While there was evidence of toxicity and a 30 to 40% decrease in biomass, the survival rate of plants was 100%. A nearly 800-fold enrichment in leaf 15 N content was observed when Fe was excluded from the nutrient solution during the treatment period, presumably due to increased plant uptake of the Fe in the ferrocyanide. While this study provided evidence suggesting ferrocyanide transport in plants, the results raised questions as to the fate of the cyanogenic nitrogen atom. The biochemical form of this nitrogen was not assessed in this study, nor was ferrocyanide solubility adequately controlled. A more detailed study, using nutrient solutions developed with the aid of computer speciation modeling, showed that diamond willow exposed to 15 N-labeled ferrocyanide both transported and

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metabolized this molecule [26]. The initial cyanide speciation and content, 2 mg l−1 ferrocyanide as CN, was maintained during the 20-day course of the experiment, with only minimal dissociation observed. Analysis of the root and leaf tissues at the end of the experiment demonstrated dramatic increases in 15 N enrichment (5500 and 250, respectively). The cyanide content of root tissues corroborated the 15 N data, demonstrating that 15 N in the root was present as ferrocyanide. Relative to the 15 N concentration, little cyanide was detected in either root or leaf tissue, indicating ferrocyanide metabolism by the plant. The metabolic rate of ferrocyanide uptake, transport, and metabolism was lower than those for free cyanide, as only 8% of the total ferrocyanide mass was removed from the solution by willow tissues over the 20-day period (with 100% cyanide recovery). A study conducted with barley, oat, and wild cane provided similar results. When these plants were exposed to 5 mg ferrocyanide l−1 using the same hydroponic system, 15 N enrichments (Figure 24.2) from 750 to 1750 and from 500 to 600 were observed in roots and leaves of these plants, respectively [31]. These grasses showed a more modest enrichment in roots, but a greater enrichment in leaves as compared to diamond willow. As with diamond willow, barley and oat showed a greater enrichment in roots. Enrichment for wild cane was similar for root and shoot tissues. Even though a detailed analysis of cyanide content and speciation was not conducted for these additional plants, the results suggest that the propensity to transport and metabolize complex cyanide compounds such as ferrocyanide differs between plant species. Varied metabolism of cyanide relative to tissue and plant developmental age has been reported earlier [35]. The presumed differences in ferrocyanide metabolism may be due to differences in developmental age between the willow, barley, oat, and wild cane plants used in these studies or may simply reflect differential distribution of metabolites bearing the 15 N atom. Studies are currently under way to identify the biochemical form of 15 N in these plant tissues. The constructed wetlands described in Section 24.3.2.1 [32] also removed iron complexed cyanide under field conditions. While the cyanide in the inflow water for the first demonstration was not specifically shown to be present as complexed cyanide compounds, the presence of Fe at >1 mg l−1 strongly favors complex formation. The cyanide concentrations in water pumped into the wetland for the first demonstration were 0.03 to 0.05 mg l−1 total cyanide. There was a statistically significant decrease in total cyanide (88%) within the wetland, with the total cyanide concentration near the midpoint of the wetland dropping below the laboratory reporting limit of 0.01 mg l−1 [32]. The wetland demonstration study at Duck Spring, Alcoa, TN, described earlier, yielded results that are more convincing as far as reduction in iron–cyanide complexes is concerned. Approximately 97% of the total cyanide that was primarily composed of iron–cyanide complexes (90%), was removed in course of the 21-day monitoring at the wetland. Analysis of these results suggests that photodissociation of iron–cyanide complexes to free cyanide, followed by biodegradation of free cyanide in the water column, is the principal mechanism (∼70%) of cyanide removal at this wetland facility. Rhizosphere-mediated biodegradation of free cyanide appears to be a secondary factor in the removal of the cyanide (accounting for ∼30%). Composite plant tissue samples analyzed for cyanide yielded average cyanide concentration of 3 and 5 mg day−1 (DW) in stems and roots, over 4 months of active treatment period.

24.4 MODELING CYANIDE UPTAKE AND FATE IN PLANTS Plant-scale process modeling can be used to augment experimental efforts by identifying which plant processes contribute to chemical transport and metabolism. Modeling can be used to target and optimize specific physiological parameters that are important to phytoremediation efforts. A model for cyanide species uptake by plants was developed [36] to interpret quantitatively the data from hydroponic uptake experiments with willow, a representative plant used in phytoremediation projects [26]. The model was developed reflecting the relevant plant physiology to gain insight into the relative

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(a) 2000

a

Barley Oat Wild cane

ab d15N ‰

1500 b

1000

500 0

(b)

[15N], mg g–1 DW

500 400

a ab

300

a

b

ab b

200 100 0

(c) Total N, mg g–1 DW

60,000

a

50,000

b

40,000 c

30,000 20,000 10,000 0 Root

Shoot Tissue

FIGURE 24.2 Impact of ferrocyanide treatment on 15 N enrichment (δ15 N‰) (a), 15 N concentration (b), and total tissue nitrogen (c) in plant roots and shoots. Data represent the mean and (±1) standard error (barley n = 8; oat n = 4; wild cane n = 5). Within a plant organ (root or shoot), plant tissues from specific plants that are assigned different letters differ significantly from plants within that same group (p ≤ .05). (Source: Reprinted from Environ. Pollut., 127, Samiotakis, M. and Ebbs, S.D., Possible evidence for transport of an iron cyanide complex by plants, 169, Copyright 2004, with permission from Elsevier.)

role of different processes in free and complexed cyanide transport and assimilation in plants, and to determine process rates under the experimental conditions. For engineering assessment and design of phytoremediation processes, models developed for field scale and hence encompassing a population of plants will be needed. Initial modeling efforts, however, have focused on cyanide uptake and fate in the single plant. Uptake of a chemical into the root of a plant depends on the properties of the chemical. Previous models [19,37–39] dealt with organic molecules that can diffuse across the root cell membrane and partition to the transpiration stream [40]. A previous study speculated that transport of ferrocyanide may require active (protein-mediated) chemical uptake [28], while another provided evidence of

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passive transport through mitochondrial channels [41]. As anions, CN− and metal-complexed dissolved cyanide species such as ferrocyanide, [Fe(CN)6 ]4− , are impeded from passive flow into root cells [40,42]. Protein-mediated uptake has been shown to be important in the movement of nutrients and ionic chemicals into plant tissue [40,43,44]. Protein-mediated uptake was included for the cyanide model using saturation kinetics, as suggested by two chemical migration models [43,45] and previous metal uptake experiments [46–48].

24.4.1 MODEL STRUCTURE The plant uptake model [36] was constructed using a compartmentalized approach with mass transfer and reaction processes governing chemical fate within the system, with the willow tree as an example plant system (Figure 24.3). Plant tissue was separated into root, stem, and leaf compartments in lieu of the data collected in hydroponic experiments [26]. In solution, cyanide concentrations and volatilization were also included as shown in Figure 24.3. Root tissue was separated further to account for protein-mediated uptake of chemical across the root endodermis and apoplastic movement [37,39]. The apoplast is a continuous pathway from the bulk solution to the endodermis that flows through plant cell walls, but not within plant cells, thereby providing a low resistance mechanism for solute movement [49]. Mass balances were developed around each compartment using the mass transfer processes for cyanide movement between, and reaction processes within, compartments. The mass balance equations were combined to form a model describing the fate of cyanide compounds in a plant–water (hydroponic) system. Advection and diffusion were assumed to govern the movement between

Air g, l

V

Leaf

7

A Stem g, l Root interior 4

Cell wall

5

Aeff Root

U

S

Interstitial space

6

3

A, D Control S losses 1

CN–

k Fe(CN)64–

Solution

2

FIGURE 24.3 Schematic of plant model for uptake and metabolism of cyanide showing major compartments. The root interior represents fraction of root inside the endodermis while the interstitial space represents the root tissue outside the endodermis. Transfer and reaction processes affecting the movement of a chemical species within a hydroponic system including solution and plant compartments. Transfer processes include advection (A), inefficient advection (Aeff ), diffusion (D), and volatilization (V ). Reactive processes include active uptake (U), reversible, unmediated ferrocyanide dissociation (k), irreversible, plant-mediated ferrocyanide dissociation (γ ), and plant-mediated free cyanide assimilation (λ). (Source: Bushey, J.T., Ph.D. thesis, Carnegie Mellon University, 2003.)

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the well-mixed compartments. Unlike previous models that dealt with organic partitioning to the transpiration stream [18,50,51], protein-mediated uptake was included for species transfer into the plant root interior. Adsorption to root tissue was included for both free cyanide and ferrocyanide species. Metabolism was assumed to occur via nonreversible, plant-mediated ferrocyanide dissociation followed by assimilation of the free form. Assimilated product was transferred within the various willow compartments in a manner similar to that for each cyanide species. The potential for free cyanide volatilization and assimilated product first-order loss from the leaf tissue was also included. In this specific system, cyanide solid precipitation, plant growth, and phloem redistribution were not included. The resulting model consisted of seven compartments and 17 parameters.

24.4.2 MODEL CALIBRATION AND SIMULATIONS Optimal values of model parameters were determined by fitting experimental data for free cyanide and ferrocyanide uptake and assimilation by willow [26], and are given in Table 24.1. Optimization results showed that free cyanide volatilization and cyanide cell wall adsorption were negligible. Free cyanide volatilization accounted for <0.01% of the initial cyanide mass within each system. The root cell wall fraction was also <0.01% for free cyanide and, while larger for ferrocyanide, was still negligible at <0.04%. Modeling results suggested that protein-mediated uptake kinetics are justified for free cyanide but not for ferrocyanide, for which the low KmFC value does not support active uptake. Examination of ferrocyanide uptake as a first-order process with the transpiration stream was also not sufficient to describe the solution concentrations. Uptake for ferrocyanide is likely to occur as a combination of the two processes [45]. The root free cyanide assimilation and

TABLE 24.1 Willow plant model [36,67] process variables determined through fitting of hydroponic uptake data for free cyanide and ferrocyanide [26] Symbol

Description

Ferrocyanide parameters α FC /β FC Ratio of cell wall adsorption/desorption eFC Transfer efficiency ff µFC max

Maximum active uptake rate constant

FC Km γRoot γLeaf

Half-saturation constant Root dissociation constant Leaf dissociation constant

Parameter Value

Units

3.10 1.11 × 10−4

l/kg µmol/µmol

535

µmol/l-day

0.503 2.85 × 10−4 1.39 × 10−5

µmol/l kg/day kg/day

Free cyanide parameters eCN Transfer efficiency ff

0.115

µmol/µmol

µCN max

Maximum active uptake rate constant

54550

µmol/l-day

CN Km λRoot λLeaf

Half-saturation constant Root assimilation constant Leaf assimilation constant

15.0 0.878 0.093

µmol/l kg/day kg/day

Assimilate parameters eA Transfer efficiency ff A HLeaf Volatilization rate constant

0.083 2.35 × 10−3

µmol/µmol kg/day

Some values constrained within possible range. Fixed plant physical parameter values and parameters eliminated from consideration, including α CN /β CN (free cyanide cell wall CN (leaf free cyanide volatilization), can be found elsewhere [36]. adsorption/desorption) and HLeaf

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ferrocyanide metabolic constants, λRoot and γRoot , were higher than those for leaf tissues, λLeaf and γLeaf . As expected, free cyanide uptake and assimilation rate constants were higher compared with ferrocyanide uptake and assimilation rate constants. The tissue concentrations predicted with the fitted model parameter values fit the data sufficiently, except for the stem and leaf tissue assimilate concentrations, particularly in the free cyanide system. The low predicted values for these compartments combined with the slightly underestimated removal of free cyanide from the KCN solution suggest a potential importance for redistribution of assimilates in the phloem. Phloem transport was not considered but could redistribute assimilated product from leaf tissue to plant root and stem tissue to reflect experimental data. Additional considerations such as plant growth effects and cyanide solid precipitation could also be added to the model. The model can be applied to systems that utilize plants other than willow, which follows calibration of the model with experimental uptake data regarding system physical properties and cyanide concentrations. Cyanide uptake, dissociation, metabolic, and volatilization rates may be plant–species specific, requiring calibration of the uptake model for each specific system application.

24.5 CYANIDE PHYTOREMEDIATION: LIMITATIONS AND CHALLENGES Laboratory and small-scale field studies have generated preliminary data that demonstrate proof of concept for phytoremediation of free- and metal-complexed cyanide in aqueous and terrestrial systems. Additional research is needed to implement phytoremediation strategies for field sites. Successful development requires that additional factors be considered, including the selection of plant species, mixed plant species vs. single species, plant spacing and density, cyanide speciation and bioavailability, regulatory concerns regarding human and ecological risk, and economics, especially the cost of land needed. Such factors represent the next challenge toward the implementation of cyanide phytoremediation at the field scale, defining and refining the boundaries of applicability.

24.5.1 SELECTION OF PLANT SPECIES As a plant-based biotechnology, the selection of plant species for use in a phytoremediation strategy is of critical importance. Plants used for phytoremediation must possess general characteristics such as high biomass, rapid growth, high transpiration rate, and the ability to grow under less-than-ideal conditions. Also necessary are characteristics specific for the contaminants of interest. In addition to representing the best possible combination of these characteristics, plant selection may be influenced by site-specific conditions or by the need to focus on “native” plant species rather than introduced species. While biotechnology has produced a variety of engineered plants with desirable traits, their open use in the field is still under regulatory scrutiny. The concern stems from the possible movement of foreign genetic material into natural populations. Until such regulatory issues are resolved, efforts for cyanide phytoremediation will have to rely on the natural cyanide assimilatory capacity of plants. 24.5.1.1 General Plant Characteristics Contaminant removal during phytoremediation is largely a function of plant biomass production [52]. This is due to several factors, including the greater root-to-shoot transport by plants with higher rates of transpiration and the dilution effect created by high biomass. This dilution effect reduces the contaminant dose throughout the plant, allowing the innate detoxification mechanisms the opportunity to remove the contaminant before tissue concentrations reach potentially detrimental levels. Moreover, the physicochemical properties of the soil and the nutrient regime at contaminated sites may support only limited plant growth. Plants species that can tolerate such conditions

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and produce nominal amounts of biomass would therefore be required, with the advantage of a larger transport of cyanide relative to intolerant plants, resulting from an increase in biomass production. Willow and poplar have been the focus of a wide variety of phytoremediation studies because these plants can grow and thrive in compromised environments, as evidenced by their presence on cyanide contaminated sites [25,28]. Both species clearly have a rapid growth rate under less-than-ideal soil conditions, as well as a tolerance to flooding. Willow and poplar can be easily cultivated through clonal propagation, which has the advantage of providing genetically identical plants. Perhaps more importantly, the high transpiration rate by these tree species creates a hydraulic lift that can facilitate contaminant removal from aqueous media. Such characteristics have stimulated interest in the use of willow and poplar species for the remediation of contaminated soil and groundwater through either in situ or biological pump-and-treat approaches. Other plants that have received attention as possible candidates for phytoremediation include Indian mustard (Brassica juncea L.), species of Amaranthus, grasses, and contaminant hyperaccumulators (plants that accumulate a given contaminant to concentrations 100-fold greater than observed in normal plants). However, these plants are generally selected because they have traits that make these species more favorable for remediation of specific contaminants, rather than contaminants in general. Indian mustard may have some utility for cyanide removal as it has been shown to assimilate selenium from the selenocyanate (SeCN− ) present in the effluent streams from oil refineries and mining wastes [53]. 24.5.1.2 Plant Characteristics Specific for Cyanide Phytoremediation Phytoremediation of cyanide requires plant species with the capacity to detoxify and assimilate cyanide at rates fast enough to prevent cyanide intoxication. Fortunately, cyanide metabolism appears to be well distributed within the plant kingdom (Chapter 6). Hence, all plants may theoretically be capable of remediating cyanide contamination. However, the fact that all plants seemingly possess the capacity to metabolize cyanide does not necessarily translate into active cyanide metabolism under experimental or field conditions, at rates sufficient to provide effective phytoremediation. The capacity to metabolize cyanide is intricately tied to the issue of cyanide toxicity discussed in Section 24.3.1. One possibility is that cyanogenic plants may have a greater potential for cyanide assimilation compared with noncyanogenic plants because their own use of cyanogenic compounds would require that pathways exist to metabolize cyanide and prevent self-intoxication. Wild cane (Sorghum bicolor L.) is a cyanogenic plant that has been shown to transport 15 N-labeled ferrocyanide [31]. The extent of the transport was similar to that of barley (Hordeum vulgare L.) and oat (Avena sativa L.). Nevertheless, the distribution of cyanogenic nitrogen (determined from measurements of 15 N in roots and shoots) revealed a different pattern than that observed in barley, oat, and willow [26]. Unlike the other three species, which retained 15 N from ferrocyanide predominantly in the root, wild cane showed an equal distribution between root and shoot. Determining whether such a pattern was due to a difference in transport and metabolism or due to some other factor will require additional study, but the results at least imply that there are differences in cyanide transport and metabolism between plant species. The fact that cyanide assimilation by plants increases during periods of physiological stress may also be an advantage for cyanide phytoremediation. The activity of the mitochondrial alternative oxidase and production of the stress hormone ethylene are both involved in plant response to stress and are both related in part to cyanide metabolism (Chapter 3). Less than ideal conditions in the field, including both exposure to contaminants (including cyanide) and any adverse environmental conditions, may enhance cyanide assimilation, resulting in the potential for greater assimilation than might be observed under ideal laboratory or greenhouse conditions. Of course, a greater rate of cyanide transport and metabolism would only be beneficial if the cyanide were bioavailable, based upon the physicochemical conditions of the media (see Section 24.5.2).

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One possible approach to selecting a species for cyanide phytoremediation would be to select species previously exposed to cyanide under less-than-optimal conditions, operating under the assumption that plant species growing under these conditions may be metabolizing the cyanide contamination, providing one reason why willow and poplar have been the most frequently studied. The clone of diamond willow used previously [26] grows in proximity to cyanide contamination in a soil with a high water table that floods frequently. The nutrient content of the soil is also poor. The occurrence of this clone in proximity to cyanide contamination implies the presence of additional, perhaps specific, characteristics that are advantageous for cyanide phytoremediation. Several other studies used similar justification in plant selection [25,33]. Such an approach would also be advantageous because of the selection of “native” plant species for phytoremediation. A possible disadvantage is the potential bioaccumulation of cyanide within plant aerial tissue as discussed in Section 24.7.1. Another important factor with respect to cyanide phytoremediation is the ability of the plant species to transport and metabolize different chemical species of cyanide. The aforementioned ubiquity of cyanide metabolism in plants may be limited to free cyanide and thiocyanate. Phytoremediation of complexed cyanides may prove more challenging as there have been only few studies demonstrating transport and metabolism of these compounds. These studies have suggested that trees, grasses, and emergent aquatic plants have the capacity for complexed cyanide phytoremediation, implying that a wide range of plants may be capable of remediating complexed cyanide compounds. Phytoremediation strategies for ameliorating cyanide contamination in both terrestrial and aqueous systems are possible. However, previous studies (see Section 24.3.2.2) illustrated the differences in the transport and metabolism that exist between free and complexed cyanide species. Removal has also been shown to be specific to the plant species utilized. To be successful, a phytoremediation strategy would require a plant species that transports and metabolizes complexed cyanide compounds at the most rapid, sustainable rate possible, and that survives under high density cultivation conditions.

24.5.2 USE OF MULTIPLE SPECIES Many phytoremediation efforts take an “agronomic” approach to environmental remediation. The emphasis is placed on a single plant species that is planted and managed much like an agricultural crop. For some species, such as Indian mustard, several crops can be grown each season. An alternate approach that might prove beneficial for cyanide phytoremediation would be to use an ecological “community” based approach. Natural cyanide cycling exists in the environment (Chapter 12), demonstrating the symbiosis between organisms within a community to work commensally and competitively to produce and metabolize cyanide. The use of multiple plant species working simultaneously, in concert with bacteria and fungi, may achieve a higher cyanide metabolism compared with that obtained by any single organism alone. Since the preferred goal of cyanide phytoremediation is phytodegradation, rather than phytoextraction, a more holistic approach may be ultimately more beneficial. The constructed wetland efforts (Sections 24.3.2.1 and 24.3.2.2) [32,54] demonstrate the use of a mixed species approach for cyanide removal. In addition to plants, micro-organisms were isolated from the sediment within the system and were shown to have the ability to metabolize free cyanide.

24.6 POSSIBLE SCENARIOS FOR CYANIDE PHYTOREMEDIATION A necessary step in assessing phytoremediation as a viable remediation option is a field-scale test. Without evidence of removal on a large scale under field conditions, which account for seasonal

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changes in weather and plant growth (including periods of plant dormancy), the true effectiveness of phytoremediation of cyanide remains in doubt. Laboratory results strongly suggest that willows are able to take up and convert free cyanide and ferrocyanide to assimilation products. Field-scale testing, particularly for complexed forms, will provide information on the cumulative effect of the removal processes. Hydroponic testing has shown potential for plant uptake, but has excluded by design important contributory processes such as precipitation, sorption, microbial degradation, photodissociation, and volatilization. These processes may contribute equally to or greater than the cyanide removal by the plants. Further, unforeseen benefits may occur as a result of the processes inherent to the natural cyanide cycle.

24.6.1 ENGINEERING IMPLICATIONS A simple engineering assessment was performed for phytoremediation using uptake data from 20-day hydroponic experiments [26] with willow. The size and planting density of an existing pilot wetland system [55] were used to determine the critical concentration that the wetland could handle. Assumptions included: a plant weight of 1 kg (fresh), 1-ft × 1-ft (0.305-m×0.305-m) plant spacing (assumed for this scoping analysis, but probably too dense for willows), total wetland area of 2.6 acres (10,522 m2 ) and a planting density of 75% of the plan area. The critical concentration, as defined by the maximum amount of cyanide entering the system without breakthrough, was calculated by a simple balance on the cyanide within the system. The analysis was performed for two scenarios, one for cyanide entering in the free form and one for cyanide entering in the complexed form. Ferrocyanide loading was separated further into predictions of the amount of input cyanide taken up, and that assimilated, by wetland plants. Free cyanide was almost completely metabolized within plant tissue. Therefore, cyanide uptake was not examined separately from that assimilated. Uptake predictions were made using hydroponic data [29] and the assumptions regarding plant size and density. While some of the assumptions, for example, the planting density of 1 kg fresh weight per ft2 (0.093 m2 ) of treatment area, are not necessarily appropriate for the willow plants used in the hydroponic study, the results could be adjusted relative to the assumed values. The simulated performance of the hypothesized wetland phytoremediation system for different influent concentrations of free cyanide and ferrocyanide is presented in Figure 24.4. The critical normalized uptake ( y-axis) value is 1.0, the point at which all of the cyanide entering the system is taken up or assimilated by the plants. A point on the respective cyanide uptake curve below 1.0 represents an unsaturated treatment system, while that greater than 1.0 represents cyanide breakthrough. The critical treatment concentration can be determined from the point on the uptake curves that produces a normalized uptake ratio of 1.0. Simulation results in Figure 24.4 show that the particular test wetland could completely assimilate a ferrocyanide input concentration less than 0.09 ppm as CN, and a free cyanide input concentration of 1.0 ppm as CN, for a hydraulic residence time (HRT) of 5.1 days through the 3.5 million liter volume. For uptake of ferrocyanide into the plant, the wetland could handle 0.2 ppm as CN. These values are in the range of water concentrations present at many contaminated sites. For a 2.6 acre wetland, this represents a loading and degradation rate of 0.84 kg CNTot ha−1 day−1 . The estimated maximum input concentrations are conservative, in that cyanide uptake by plants is only one of the many loss processes for cyanide within a wetland system, as depicted in Figure 12.6 in Chapter 12. Biodegradation, volatilization, adsorption, absorption, and precipitation are additional removal processes that potentially can affect cyanide within a wetland treatment system. Another important process in the removal of ferrocyanide, particularly in surface systems, is photodissociation to the much more reactive free form (see Chapter 9). If photodissociation were considered, the potential uptake would be similar to that for the free cyanide uptake. The potential ferrocyanide removal would be limited by free cyanide plant uptake, not photodissociation. System design modifications could improve removal efficiency. For example, the flow loadings to the wetland could be reduced to provide more cyanide removal from the system, but this would

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Total CN uptake/assimilation by plants/ Total CN entering system

2.0 Ferrocyanide uptake Ferrocyanide assimilation Free cyanide uptake/assimilation Limit for process effectiveness = 1.0

1.5

1.0

0.5

0.0 0.0

0.1

0.2

0.3

0.9

Inflow concentration (mg

l–1

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as CN)

FIGURE 24.4 Simulated performance of a hypothetical wetland phytoremediation system for different assumed influent concentrations of free cyanide and ferrocyanide. Model simulations based upon data from hydroponic uptake experiments with willows. The y-axis shows the total amount of cyanide taken up or assimilated by the plants normalized by the total amount of cyanide entering the system. The critical y-axis value is 1.0 — the point at which the plants take up all the cyanide mass that enters the system. The uptake was calculated from the total amount accumulated in plant tissue over the 20-day experiment with the critical concentration based on a system mass balance. Assumptions include (1) 1-ft×1-ft (0.305-m×0.305-m) plant spacing; (2) 75% plan area surface coverage; and (3) each plant mass is 1 kg fresh weight. (Source: Bushey, J.T., Ph.D. thesis, Carnegie Mellon University, 2003.)

increase land requirements. As previously discussed, the planting density of the system can also be adjusted, but there is a limit to planting density that varies with plant species. One limitation to the system is the assumption of continual treatment. Seasonal dependencies can be accounted for by scaling the absolute mass of cyanide removed by the system with the fraction of time the system is active for. During winter dormancy, it is likely that the abiotic loss processes (i.e., adsorption and solid precipitation) would dominate any removal of cyanide from the aqueous phase.

24.6.2 FIELD APPLICATIONS Field application for phytoremediation looks promising based on the Alcoa pilot study [32] and the results of the laboratory hydroponic study [26]. At the Alcoa facility, iron cyanide contaminated groundwater has surfaced through a natural spring known as Duck Spring [54]. A pilot-scale open surface wetland is in operation and has been treating ∼10 gpm flow of cyanide-contaminated spring water, as described earlier in this chapter. The promise of phytoremediation for treatment of cyanide species has also been demonstrated in field tests with wetland microcosms used to treat coal gasification wastewater that contains selenocyanate (SeCN− ) [56,57]. Fourteen different plant species, including cattail, thalia, and rabbitfoot grass, were tested in the microcosms. Significant removal of selenium mass (79%) was achieved. Accumulation in plant tissues accounted for less than 5% of the selenium removal, with incorporation in sediments accounting for most of the mass removal.

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24.7 REGULATORY CONCERNS A principal regulatory concern for environmental contaminants is their fate, particularly with respect to human or ecological impacts. Regulatory acceptance of a remediation technology requires that the approach not exacerbate increasing human or ecological risk, or promote the release of the contaminant to the environment. Two of the primary concerns associated with cyanide phytoremediation, bioaccumulation and volatilization, relate to these possibilities.

24.7.1 CYANIDE BIOACCUMULATION IN PLANT TISSUE Cyanogenic organisms, including plants and insects, use cyanide as a deterrent to herbivory. Wildlife can distinguish cyanogenic from noncyanogenic vegetation and preferentially select noncyanogenic tissues. Coevolution has also resulted in herbivores capable of consuming cyanogenic vegetation without adverse effects. Several insects have shown resistance to the cyanogenic properties of plant tissues [58], and even sequester the compounds produced by plants as a defensive device. One such example involves the larvae of the five spot burned moth (Zygaena trifolii L.), which feeds on the leaves of the cyanogenic bird’s foot trefoil (Lotus corniculatus L.) [59], in which both plant and insect contain that same array of cyanogenic compounds. The greater concern with respect to cyanide phytoremediation is that cyanide would bioaccumulate in the tissues of the plants used, though available information indicates that cyanide compounds do not bioaccumulate. Local wildlife could consume that plant tissue and be inadvertently exposed to cyanide. For example, in the constructed wetland study [32], snapping turtles migrated into the wetland and used the planted bulrush (Scirpus validus L.) as a food source. The turtles were safely and quickly relocated as a safety precaution. Such occurrences demonstrate the concern about possible ecological impacts of cyanide phytoremediation efforts. Establishing vegetation, whether in a terrestrial or aqueous system, presents wildlife a potential food source that was not previously present. In terrestrial settings where fertilizers or other management techniques are utilized to increase growth, this vegetation may be more appealing that other food sources. Insuring that cyanide bioaccumulation does not present an ecological risk in such cases is a high priority. Some studies with willow and other tree species have reported increased concentrations of cyanide in aerial tissues following exposure to free cyanide in solution culture [33]. Other studies conducted in solution culture have provided evidence to the contrary [26], showing little difference in cyanide content between treated and control plants. Likewise, willow and elderberry growing on a cyanidecontaminated gas works site showed increased plant tissue cyanide concentrations only in roots [25]. The cyanide concentration in leaves and fruits was generally lower than that observed in control plants. Field observation is one reason why bioaccumulation is not likely to represent a limiting factor for cyanide phytoremediation. Previous research has shown that plants under stress concomitantly produce the gaseous hormone ethylene and cyanide gas (Chapter 3). In addition, the presence of ethylene has been shown to increase activity of the enzymes responsible for cyanide assimilation [60–63]. At higher cyanide concentrations, the stress created by cyanide exposure would trigger ethylene release that would as well increase the rate of in vivo cyanide metabolism. Cyanide is a potent metabolic inhibitor, potentially affecting mitochondrial respiration, photosynthesis, glucose metabolism, and various enzymes (e.g., carbonic anhydrase, alkaline phosphatase). The mitochondrial alternative oxidase (Chapter 6) would act to minimize the impact of cyanide on respiratory metabolism, but not for prolonged periods of time, due to the energy limitations. The likelihood that cyanide eventually would inhibit critical processes makes bioaccumulation of cyanide in plant tissues unlikely. Cyanide-exposed plants would either die or survive the cyanide exposure by metabolizing the cyanide rapidly enough to prevent intoxication. Proper selection of a plant species for use in specific phytoremediation scenarios would possibly avoid plants that were either sensitive to cyanide toxicity or tended to bioaccumulate cyanide

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compounds. A simple precaution would be to conduct a simple series of preliminary experiments, as is done in most phytoremediation studies, to show that toxicity and bioaccumulation are not issues. The assimilatory pathways for cyanide involve conversion to amino acids, not cyanide-releasing compounds such as cyanogenic glycosides. Synthesis of cyanogenic glycosides forms the cyanide moiety after an amino group has been conjugated to the sugar, rather than involving conjugation of a cyanide molecule to the sugar. Additionally, cyanide represents a valuable source of carbon and nitrogen that is more effectively utilized by the plant than when simply stored, particularly since the compound being stored is potentially toxic.

24.7.2 CYANIDE VOLATILIZATION FROM PLANT TISSUES Perhaps the most significant regulatory issue associated with cyanide phytoremediation is the prospect of cyanide volatilization. Gaseous cyanide is released by nearly all plants at low levels during synthesis of ethylene or from cyanogenic plants during periods of herbivory or stress. Activity of the assimilatory pathways provides a significant degree of homeostasis, maintaining cyanide levels at concentrations below those that would poison the plant [64]. The concern is that during phytoremediation, cyanide may be transported to foliar tissues at a rate greater than the rate of metabolism. The excess cyanide may be volatilized to the atmosphere biologically in an effort to prevent cyanide toxicity, or may volatilize naturally from the leaves via physical processes. Plants would extract cyanide from the media but would subsequently release that cyanide to the atmosphere as HCN(g) . For field-scale phytoremediation efforts, volatilization could create regions of elevated gaseous cyanide concentrations that could have adverse effects on local flora and fauna. Similar regulatory concerns were raised over the prospect of mercury (Hg) phytovolatilization [16], seriously undermining attempts to develop Hg phytoremediation. For cyanide phytoremediation to be successful, the potential for cyanide phytovolatilization must be addressed and shown not to be a significant route for cyanide transport. The preliminary studies of cyanide phytoremediation that have been conducted have produced some evidence of cyanide volatilization by plants [28,33]. Willow (Salix sp.) plants exposed to free cyanide reportedly volatilized cyanide [25] at a rate dependent upon light exposure and the transpiration rate. As light exposure opens stomates and transpiration can only occur via open stomates, the relationships were suggestive of volatilization through the leaf stomata and imply that volatilization may be mediated simply by physical processes rather than a specific biological process. Support is provided by the fact that cyanide volatilization represented only a small fraction of the total cyanide mass taken up by the plants. One interpretation of these results is that cyanide volatilization may only occur under certain conditions, such as when cyanide is first transported to leaves via the transpiration stream or when cyanide levels in leaves increase transiently higher than the rate of assimilation. For HCN release from the leaves, these events would most likely have to be coincident with stomatal opening. As the mass fraction of cyanide lost via leaf tissue in the mass balance is relatively small, volatilization in this situation would appear not to be a sustained phenomenon. No volatilization of cyanide was observed for Indian mustard or muskgrass exposed to selenocyanate [53], or diamond willow (S. eriocephala L.) or pea (Pisum sativum L.) [65] exposed to free cyanide. A mass balance calculated for diamond willows exposed to free cyanide achieved only a 60% recovery [26]. Although preliminary experiments were negative, volatilization of HCN from leaf tissue may have accounted for the remainder of the cyanide. The mass balance in this case was calculated using chemical analyses for cyanide in solution and stable isotope measurements for tissue, following exposure to 15 N-labeled potassium cyanide. Neither analyses accounted for volatile cyanide per se, but indicated a loss of initial 15 N from the system. While these losses could represent cyanide volatilization, another possibility is that other volatile compounds released by willow may have contained the 15 N atoms from the labeled cyanide compounds. Subsequent biochemical analyses of leaves from willows demonstrated extensive assimilation of 15 N from potassium cyanide into primary metabolism. Trees such as cottonwood (Populus deltoides L.) release a variety of

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volatile compounds, including nitrogen-containing compounds such as dimethylallyl phosphate [66]. Assimilation of cyanide could have resulted in the incorporation of 15 N into volatile compounds that were released from leaves rather than the parent cyanide compound. The plant-scale phytoremediation process model (Section 24.4) predictions for fate and transport of cyanide species in willow are in agreement with the study [25], suggesting a minor role for volatilization. However, further study of cyanide volatilization by plants is clearly warranted to validate this prediction. Such studies should be conducted under controlled environmental conditions or, if possible, field conditions to assess the role of environmental factors on this process. Efforts were undertaken to determine whether cyanide was volatilized from the constructed wetland described above [32], but there was only one instance where cyanide was detected. This may demonstrate that the monitoring of cyanide volatilization may require sensitive assays. Cyanide volatilization can be assessed through a variety of techniques, including Draeger tubes and alkaline picrate papers. Draeger tubes have a functional range of 2 to 15 mg m−3 . Alkaline picrate papers have been used largely to monitor cyanide emissions from cyanogenic plants, but are subject to measurement interpretation. Whether these techniques have the sensitivity to measure cyanide volatilization from tissues of noncyanogenic plants exposed to cyanide remains to be seen.

24.8 SUMMARY AND CONCLUSIONS • Phytoremediation is a family of biotechnologies that use plants for the remediation of environmental contamination. Phytoremediation has been applied to numerous contaminants, including metals and metalloids, uranium and fission products, hydrocarbons, BTEX compounds, petroleum products, ammunition wastes, chlorinated solvents, and herbicides. Cyanide phytoremediation is a new area that is receiving increasing attention. • Cyanide is one of the most feasible targets for phytoremediation because all plants appear to possess metabolic pathways that metabolize cyanide. These pathways are present because cyanide is a common constituent of plant metabolism. Evidence suggests that these pathways can also metabolize exogenous cyanide. • While most studies of cyanide phytoremediation have been conducted in hydroponic culture, a few studies have been conducted in the field. These preliminary studies have all supported the technical prospect of cyanide phytoremediation, demonstrating removal, and, in some cases, metabolism of cyanide contamination. The data from these studies have laid the foundation for future studies of cyanide phytoremediation as well as for the development of scenarios for the deployment of phytoremediation in the field. • Modeling of cyanide phytoremediation using hydroponic data has suggested that volatilization of free cyanide is negligible and that active uptake kinetics may only be appropriate for the free form but not for complexes. The inclusion of redistribution, particularly for assimilates, could improve the model to make it fit for the data. However, modeling remains an important tool to assist in understanding the potential for field implementation of cyanide phytoremediation. • The selection of plants for use in cyanide phytoremediation requires plants with a specific suite of traits. General traits of benefit for cyanide phytoremediation include high biomass, rapid growth, high transpiration rate, and the ability to grow under less-than-ideal conditions. Traits specific for cyanide that would benefit phytoremediation include rapid rates of cyanide metabolism and the ability to transport and metabolize a range of cyanide compounds. • Treatment scenarios have been limited to a few, selected wetland systems that have demonstrated the potential for remediating free and complexed cyanide through phytoremediation applications. The effectiveness of the treatment systems is still being evaluated as a more complete understanding of cyanide fate is necessary.

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• Although some studies have reported an increase in the concentrations of cyanide in plant tissues following exposure to cyanide in the external media, cyanide bioaccumulation would be unlikely during phytoremediation. The innate pathways for cyanide metabolism in plants would typically act to maintain cyanide concentrations at the lowest level possible, to prevent cyanide intoxication. Should these pathways fail to maintain cyanide homeostasis, the plants would die. Plant species that do bioaccumulate cyanide (if any) would in all likelihood be excluded from consideration for phytoremediation and would not be deployed in the field. • The potential volatilization of cyanide by plants represents the most significant obstacle to regulatory acceptance of cyanide phytoremediation. Several studies have reported possible volatilization of cyanide by plants exposed to cyanide, but most have not provided conclusive evidence. Additional studies using highly sensitive equipment or assays will likely be required to evaluate the potential for cyanide volatilization.

REFERENCES 1. McCutcheon, S.C. and Schnoor, J.L., Overview of phytotransformation and control of wastes, in Phytoremediation: Transformation and Control of Contaminants, McCutcheon, S.C. and Schnoor, J.L., Eds., Wiley-Interscience, New York, 2003, p. 3. 2. Cunningham, S.D., Berti, W.R., and Huang, J.W., Phytoremediation of contaminated soils, Trends Biotechnol., 13, 393, 1995. 3. Dushenov, S., Nanda Kumar, P.B.A., Motto, H., and Raskin, I., Rhizofiltration: the use of plants to remove heavy metals from aqueous streams, Environ. Sci. Technol., 29, 1239, 1995. 4. Dushenov, S., Vasudev, D., Kapulnik, Y., Gleba, D., Fleisher, D., Ting, K.C., and Ensley, B., Removal of uranium from water using terrestrial plants, Environ. Sci. Technol., 29, 1239, 1995. 5. Carapeto, C. and Purchase, D., Distribution and removal of cadmium and lead in a constructed wetland receiving urban runoff, Bull. Environ. Contam. Toxicol., 65, 322, 2000. 6. Lemly, A.D. and Ohlendorf, H.M., Regulatory implications of using constructed wetlands to treat selenium-laden wastewater, Ecotoxicol. Environ. Saf., 52, 46, 2002. 7. Lin, Z.Q. and Terry, N., Selenium removal by constructed wetlands: quantitative importance of biological volatilization in the treatment of selenium-laden agricultural drainage water, Environ. Eng. Sci., 37, 606, 2003. 8. Giraud, F., Guiraud, P., Kadri, M., Blake, G., and Steiman, R., Biodegradation of anthracene and fluoranthene by fungi isolated from an experimental constructed wetland for wastewater treatment, Water Res., 35, 4126, 2001. 9. Higgins, J. and Maclean, M., The use of a very large constructed sub-surface flow wetland to treat glycol-contaminated stormwater from aircraft-de-icing operations, Water Qual. Res. J. Canada, 37, 785, 2002. 10. McCutcheon, S.C. and Rock, S.A., Phytoremediation: state of the science conference and other developments, Int. J. Phytoremediation, 3, 1, 2001. 11. Schwitzguebel, J.P., Aubert, S., Grosse, W., and Laturnus, F., Sulphonated aromatic pollutants: limits of microbial degradability and potential of phytoremediation, Environ. Sci. Pollut. Res. Int., 9, 62, 2002. 12. Lytle, C.M., Lytle, F.W., Yang, N., Qian, J.H., Hansen, D.J., Zayed, A., and Terry, N., Reduction of Cr(VI) to Cr(III) by wetland plants: potential for in situ heavy metal detoxification, Environ. Sci. Technol., 32, 3087, 1998. 13. Salt, D.E., Blaylock, M., Kumar, N.P.B.A., Dushenkov, V., Ensley, B.D., Chet, I., and Raskin, I., Phytoremediation: a novel strategy for the removal of toxic metals from the environment using plants, Biotechnology, 13, 468, 1995. 14. de Souza, M.P., Pilon-Smits, E.A.H., Lytle, C.M., Hwang, S., Tai, J., Honma, T.S.U., Yeh, L., and Terry, N., Rate-limiting steps in selenium assimilation and volatilization by Indian mustard, Plant Physiol., 117, 1487, 1998. 15. Pilon-Smits, E.A.H., De, S.M.P., Lytle, C.M., Shang, C., Lugo, T., and Terry, N., Selenium volatilization and assimilation by hybrid poplar (Populus tremula X alba), J. Exp. Bot., 49, 1889, 1998.

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16. Meagher, R.B., Engineered phytoremediation of mercury pollution in soil and water using bacterial genes, in Phytoremediation of Contaminated Soil and Water, Terry, N. and Banuelos, G.S., Eds., Lewis, Boca Raton, FL, 2000, p. 201. 17. Orchard, B.J., Douchette, W.J., Chard, J.K., and Bugbee, B., Uptake of trichloroethylene by hybrid poplar trees grown hydroponically in flow-through plant growth chambers, Environ. Toxicol. Chem., 19, 895, 2000. 18. Burken, J.G. and Schnoor, J.L., Predictive relationships for uptake of organic contaminants by hybrid poplar trees, Environ. Sci. Technol., 32, 3379, 1998. 19. Burken, J.G. and Schnoor, J.L., Uptake and metabolism of atrazine by poplar trees, Environ. Sci. Technol., 31, 1399, 1997. 20. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 21. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 22. Grossman, K., Quinclorac belongs to a new class of highly selective auxin herbicides, Weed Sci., 46, 707, 1998. 23. McMahon-Smith, J. and Arteca, R.N., Molecular control of ethylene production by cyanide in Arabidopsis thaliana, Physiol. Plant, 109, 180, 2000. 24. Wallace, A., Cha, J.W., and Mueller, R.T., Cyanide effects on transport of trace materials in plants [kidney beans, phytotoxicity], Commun. Soil Sci. Plant Anal., 8, 709, 1977. 25. Trapp, S., Koch, I., and Christiansen, H., Aufnahme von Cyanid in Pflanzen: risiko oder Chance fuer die Phytoremediation? Unweldt Shcad Forsch, 13, 20, 2001. 26. Ebbs, S.D., Bushey, J.T., Poston, S., Kosma, D., Samiotakis, M., and Dzombak, D.A., Transport and metabolism of free cyanide and iron cyanide complexes by willow, Plant Cell Environ., 26, 1467, 2003. 27. Bissey, R. and Butler, O., Effect of applications of sodium chlorate and ammonium thiocyanate on subsequent sowings of wheat, Agron. J., 26, 838, 1934. 28. Reeves, M., Treatment of fluoride and iron cyanides using willow: a greenhouse feasibility study. M.S. thesis, Cornell University, Ithaca, NY, 2000. 29. Samiotakis, M. Uptake and assimilation of iron cyanide by different plant species. M.S. thesis, Southern Illinois University Carbondale, Carbondale, IL, 2002. 30. Kreitinger, J.P., Weinstein, L.H., and King, P., unpublished data, 1998. 31. Samiotakis, M. and Ebbs, S.D., Possible evidence for transport of an iron cyanide complex by plants, Environ. Pollut., 127, 169, 2004. 32. BHE Environmental, I., ALCOA Blue Pond demonstration treatment wetland final report and preliminary design, Report 990.003/004, ALCOA Tennessee Operations, Alcoa, TN, 2000. 33. Trapp, S., Larsen, M., and Christiansen, H., Experimente zum Verbleib von Cyanid nach Aufnahme in Pflanzen, Umweldt Schad Forsch, 13, 29, 2001. 34. RETEC, Duck Spring pilot scale wetland monitoring project, prepared for Alcoa Remediation Operations, Interim Annual Report, RETEC Group, Pittsburgh, PA, 2004. 35. Wurtele, E.S., Nikolau, B.J., and Conn, E.E., Subcellular and developmental distribution of β-cyanolanine synthase in barley leaves, Plant Physiol., 78, 285, 1985. 36. Bushey, J.T., Ebbs, S.D., and Dzombak, D.A., Model for cyanide uptake by willow: model development, Int. J. Phytoremediation, in press, 2006. 37. Lindstrom, F.T., Boersma, L., and McFarlane, J.C., Mathematical model of plant uptake and translocation of organic chemicals: development of the model, J. Environ. Qual., 20, 129, 1991. 38. Trapp, S. and McFarlane, J.C., Plant Contamination: Modeling and Simulation of Organic Chemical Processes, Lewis Publishers, Boca Raton, FL, 1995. 39. Trapp, S., McFarlane, J.C., and Mathies, M., Model for uptake of xenobiotics into plants: validation with bromacil experiments, Environ. Toxicol. Chem., 13, 413, 1994. 40. Marschner, H., The Mineral Nutrition of Higher Plants, Academic Press, London, 1986. 41. Beavis, A.D. and Vercesi, A.E., Anion uniport in plant mitochondria is mediated by a Mg2+ -insensitive inner membrane anion channel, J. Biol. Chem., 267, 3079, 1992. 42. Ryan, P.R., Delhaize, E., and Jones, D.L., Function and mechanism of organic anion exudation from plants roots, Ann. Rev. Plant Physiol. Plant Molec. Biol., 52, 527, 2001.

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43. Ben-Asher, J., Simplified model of integrated water and solute uptake by salts- and seleniumaccumulating plants, Soil Sci. Soc. Am. J., 58, 1012, 1994. 44. Nye, P.H. and Tinker, P.B., Solute Movement in the Soil–Root System, University of California Press, Berkeley, CA, 1977. 45. Somma, F., Hopmans, J.W., and Clausnitzer, V., Transient three-dimensional modeling of soil water and solute transport with simultaneous root growth, root water and nutrient uptake, Plant Soil, 202, 281, 1998. 46. Cohen, C.K., Fox, T.C., Garvin, D.F., and Kochian, L.V., The role of iron-deficiency stress responses in stimulating heavy metal transport in plants, Plant Physiol., 116, 1063, 1998. 47. Hart, J.J., Norvell, W.A., Welch, R.M., Sullivan, L.A., and Kochian, L.V., Characterization of zinc uptake, binding and translocation in intact seedlings of bread and durum wheat cultivars, Plant Physiol., 118, 219, 1998. 48. Lasat, M.M., Pence, N.S., Garvin, D.F., Ebbs, S.D., and Kochian, L.V., Molecular physiology of zinc transport in the Zn hyperaccumulator Thlaspi caerulescens, J. Exp. Bot., 51, 71, 2000. 49. Nobel, P.S., Physiochemical and Environment Plant Physiology, 2nd ed., Academic Press, New York, 1999. 50. Briggs, G.G., Rigitano, R.L.O., and Bromilow, R.H., Physico-chemical factors affecting uptake by roots and translocation to shoots of weak acids in barley, Pestic. Sci., 19, 101, 1987. 51. Schnoor, J.L., Phytoremediation: technology evaluation report, TE-98–01, Ground-Water Remediation Technologies Analysis Center, Pittsburgh, PA, 1997. 52. Ebbs, S.D., Lasat, M.M., Brady, D.J., Cornish, J., Gordon, R., and Kochian, L.V., Phytoextraction of cadmium and zinc from a contaminated soil, J. Environ. Qual., 26, 1424, 1997. 53. de Souza, M.P., Pickering, I.J., Walla, M., and Terry, N., Selenium assimilation and volatization from selenocyanate-treated Indian mustard and muskgrass, Plant Physiol., 128, 625, 2002. 54. GTI, Development of environmental acceptable endpoints for impacted sediments, and groundwater at MGP sites, GTI Contract/Report No. 8643, Gas Technology Institute, Des Plaines, IL, 2004. 55. USEPA, Constructed wetlands treatment of municipal wastewaters, EPA/625/R-99/010, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 2000. 56. Terry, N., Sambukumar, S.U., and LeDuk, D.L., Biotechnological approaches for enhancing phytoremediation of heavy metals and metalloids, Acta Biotechnol., 23, 281, 2003. 57. Ye, Z.H., Lin, Z.Q., Whiting, S.N., deSouza, M.P., and Terry, N., Possible use of constructed wetland to remove selenocyanate, arsenic and boron from electricity utility wastewater, Chemosphere, 52, 1571, 2003. 58. Engler, H.S., Spencer, K.C., and Gilbert, L.E., Preventing cyanide release from leaves, Nature, 406, 144, 2000. 59. Jones, D.A., Selective eating of the acyanogenic form of the plant Lotus corniculatus by various animals, Nature, 193, 847, 1962. 60. Hasegawa, R., Maruyama, A., Sasaki, H., Tada, T., and Esashi, Y., Possible involvement of ethyleneactivated beta-cyanoalanine synthase in the regulation of cocklebur seed germination, J. Exp. Bot., 46, 551, 1995. 61. Hasegawa, R., Tomoko, T., Yuchiro, T., and Yohji, E., Presence of β-cyanoalanine synthase in unimbibed dry seeds and its activation by ethylene during pregermination, Physiol. Plant, 91, 141, 1994. 62. Maruyama, A., Yoshiyama, M., Adachi, Y., Nanba, H., Hasegawa, R., and Esashi, Y., Possible participation of beta-cyanoalanine synthase in increasing the amino acid pool of cocklebur seeds in response to ehtylene during the pregermination period, Aust. J. Plant Phys., 24, 751, 1997. 63. Wen, J.Q., Huang, F., Liang, W.S., and Liang, H.G., Increase of HCN and beta-cyanoalanine synthase activity during aging of potato tuber slices, Plant Sci., 125, 147, 1997. 64. Yip, W.K. and Yang, S.F., Ethylene biosynthesis in relation to cyanide metabolism, Bot. Bull Acad. Sin. Taipei, 39, 1, 1998. 65. Ebbs, S.D., unpublished data, 2001. 66. Rosentiel, T.N., Fisher, A.J., Fall, R., and Monson, R.K., Differential accumulation of dimethylallyl phosphate in leaves and needles of isoprene- and methylbutenol-emitting and non-emitting species, Plant Physiol., 129, 1276, 2002. 67. Bushey, J.T., Modeling cyanide uptake by willow for phytoremediation. Ph.D. thesis, Carnegie Mellon University, Pittsburgh, PA, 2003.

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of Cyanide 25 Management in Municipal Wastewaters David A. Dzombak, Anping Zheng, Michael C. Kavanaugh, Todd L. Anderson, Rula A. Deeb, and George M. Wong-Chong CONTENTS 25.1 25.2 25.3

Sources of Cyanide in Municipal Wastewater Treatment Plant Influent . . . . . . . . . . . . . . . . . Fate of Cyanide in Municipal Wastewater Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Generation of Cyanide in Municipal Wastewater Treatment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.3.1 In-Process Cyanide Generation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.3.2 Cyanide in Scrubber Waters from On-Site Biosolids Incineration . . . . . . . . . . . . . 25.4 Management Strategies for Low-Level Cyanide in Discharges from POTWs . . . . . . . . . . . 25.4.1 Reduction of Cyanide or Cyanide Precursors. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.4.2 Request a Site-Specific Discharge Permit. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.4.3 Change Point of Compliance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.4.4 In-Plant Modification of Oxidation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.4.5 Alternative Analytical Methods for Compliance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25.5 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

502 504 507 507 509 510 511 511 512 512 512 512 513

Cyanide present in municipal sewage is derived from industrial sources. Cyanide can also be input to or generated within municipal wastewater treatment processes, such as by the introduction of biosolids incineration scrubber water or in the oxidizing conditions of the disinfection process. Cyanide compounds are commonly used in a number of industries, and are present in associated process wastewaters. Industries that generate large quantities of cyanide-bearing wastewaters include electroplating and metal finishing, chemical production, steel production, and petroleum refining (see Chapters 4 and 26). As cyanide can be toxic to the microorganisms used in biological treatment processes if present at sufficiently high concentrations, and because of concern about potential for HCN gas releases in sewers, many industrial generators of cyanide-bearing wastewater are required to pretreat their wastewater prior to discharge to publicly-owned treatment works (POTWs), for example, iron and steel manufacturers [1]. Thus, concentrations of cyanide in influent to POTWs are generally low, that is, nondetectable to 500 µg/l total cyanide [2–4]. The fate of cyanide in a municipal wastewater treatment plant depends on the form and amount of cyanide. Cyanide in raw wastewater is predominantly (e.g., 60 to 70%) in the complexed form. Free cyanide species, HCN and CN− , are readily biodegradable if not present at concentrations inhibitory to aerobic bacteria. They can also be removed from solution by adsorption onto organic particles (see Chapter 5) in primary treatment and biosolids in secondary treatment. HCN, which is relatively volatile, can be removed (“stripped”) by aeration processes within wastewater treatment plants. The understanding 501

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of the fate of metal–cyanide complexes is less well developed. Weakly associated metal–cyanide complexes are subject to decomposition and hence ultimately the same fate processes relevant to free cyanide. On the other hand, strong metal–cyanide complexes are essentially nonbiodegradable at the typical residence time of a municipal wastewater treatment plant, and they will often pass through biological treatment processes. Strong metal–cyanide complexes may be removed through adsorption on mineral particles (Chapter 5) in primary treatment. It has been observed operationally and confirmed through research [5] that cyanide can be produced in 5 to 20 ppb level quantities or greater in municipal wastewater treatment, especially in the disinfection process, which involves the use of strong oxidants such as chlorine. This is a significant challenge for POTWs, as the cyanide is generated just prior to discharge, and because the POTWs generally cannot simply eliminate, or reduce disinfection levels because of cyanide concerns as they must maintain effluent disinfection performance (e.g., to meet coliform effluent limits). Effluent limits for cyanide can be in the range of 1 to 20 ppb, when dictated by ambient water quality criteria (Chapter 18). This chapter focuses on the management of cyanide in municipal wastewaters, from industrial source control to process optimization for control of final effluent quality. It examines the sources and characteristics of cyanide in POTW influent, the removal of different forms of cyanide in the common unit processes employed in POTWs, the problem of cyanide generation in POTWs, and management strategies for low-level cyanide concentrations in POTW discharges.

25.1 SOURCES OF CYANIDE IN MUNICIPAL WASTEWATER TREATMENT PLANT INFLUENT Cyanide in influent to municipal wastewater treatment plants derives predominantly from industrial sources [3], though compounds measurable as total cyanide are usually absent from domestic wastewater as a result of pretreatment at industrial sources. Iron-cyanide compounds are used in some road salts as an anticaking agent, and thus will sometimes be present in wastewater that includes storm runoff in areas that experience winter and use such material on roadways [6]. For the most part, however, any cyanide present in municipal wastewater treatment plant influent is derived from industrial sources. The primary sources of industrial wastewater discharges to public sewer systems are electroplating and metal finishing operations, chemical production, iron and steel production, and petroleum refining [2]. Concentration ranges for cyanide species in process wastewaters from examples of these industrial sources are given in Table 25.1. As indicated in the descriptions of the sources, industrial wastewaters typically undergo various kinds of treatment prior to discharge. The last two entries in Table 25.1, for plating rinse water and salt pile runoff, are characteristics for untreated waters. Plating rinse water will usually be pretreated prior to discharge to a public sewer. Table 25.1 shows that cyanide concentrations (measured as total cyanide) in industrial wastewaters commonly discharged to POTWs are generally less than 10 mg/l and free cyanide concentrations are generally less, and often much less, than 1 mg/l. With dilution of the industrial discharges in the overall wastewater flow, influent concentrations of total cyanide at POTWs are generally less than 0.5 mg/l [4] and this cyanide is primarily in dissolved form [3]. Comparison of the free cyanide and total cyanide concentrations in Table 25.1 indicates that much of the cyanide is present as metal–cyanide complexes. This predominance of complexed cyanide is maintained during transport to the POTW and is observed in POTW influent. Lordi et al. [2] measured different forms of cyanide in influents to four POTWs that receive industrial inputs; complexed cyanide was determined to constitute 64 to 77% of the total cyanide in the influent, ranging from 0.13 to 0.40 mg/l. A similar assessment of influents to three POTWs with no industrial inputs yielded lower total cyanide concentrations (<0.076 mg/l), but a higher fraction (72 to 83%) of complexed cyanide [2].

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TABLE 25.1 Concentration Ranges for Cyanide Species in Industrial Wastewaters Industrial source Coke plant ammonia still effluent Coke plant ammonia still effluent, after biological treatment Coke plant ammonia still effluent Coke plant ammonia still effluent, after biological treatment Blast furnace gas scrubber water blowdown Chemical plant wastewater: 1 Chemical plant wastewater: 2 Coke plant wastewater Oil refinery wastewater Oil refinery wastewater Oil refinery sour water stripper effluent Oil refinery sour water stripper effluent, after biological treatment Plating company wastewater: 1 Plating company wastewater: 2 Plating rinse Road salt dock runoff

Free CN (mg/l)

Total CN (mg/l)

SCN (mg/l as CN)

NRa

3.2–4.4

280–554

Luthy and Jones [43]

NR

2.6–4.5

2.6–36

Luthy and Jones [43]

8

14

200

Ganczarczyk [44]

<0.5 <0.1

3.5

<1 <1

NR

<0.2

1.3

Ganczarczyk [44] Wong-Chong and Caruso [22] Luthy et al. [45]

NDb

0.400

50.5

Kelada [25]

0.030c

0.028

10.4

Kelada [25]

0.300c ND NR 0.25–1.7

2.100 0.010 0.1–5 1–5.1

23.6 2.24 NR 2–16

Kelada [25] Kelada [25] Thiem and Alkhatib [46] Urban et al. [47]

0.09–0.15

0.1–0.7

ND

Urban et al. [47]

0.070c

0.250

0.02

Kelada [25]

0.030c

0.030

ND

Kelada [25]

NR 2.900c

14–256 (avg = 56) 25.660

NR ND

Palmer et al. [48] Kelada [25]

Reference

a NR = not reported. b ND = not detected. c Weak acid dissociable cyanide; may include some weak metal–cyanide complexes in addition to free

cyanide.

Water pollution control regulations in the United States and other developed nations require pretreatment of various industrial dischargers for cyanide prior to discharge to POTWs. Pretreatment regulations traditionally have specified maximum allowable concentrations in industrial discharges, for example, 5 mg/l for total cyanide or less; see specific examples in Table 25.2. Regulations are evolving, however, to focus on total mass loading rather than on concentration. This allows more flexibility for complying with pretreatment objectives. The new regulatory approach, adopted by the United States Environmental Protection Agency (USEPA) [7] provides for (a) use of equivalent mass limits in lieu of concentration limits, which can be useful if a facility has instituted water conservation measures and (b) equivalent concentration limits in lieu of flow-based standards, which can be helpful for facilities that have highly variable wastewater flows. The equivalent mass limits are given in terms of weight of pollutant per weight of product. For example, the pretreatment standards

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TABLE 25.2 Examples of Industrial Pretreatment Requirements for Cyanide Imposed by Municipal Wastewater Treatment Authorities Municipal wastewater Treatment authority

Total cyanide limit for discharge to POTW (mg/l)a

Albuquerque, New Mexico Boston, Massachusetts (Massachusetts Water Resources Authority) Chicago, Illinois (Metropolitan Water Reclamation District of Greater Chicago) Flagstaff, Arizona Fresno, California Phoenix, Arizona Pittsburgh, Pennsylvania (Allegheny County Sanitary Authority) Sacramento, California (Sacramento Regional County Sanitation District)

0.45 or 0.10b 0.5

5.0

0.240 0.77 2.0 3.6 1.2

a Limits as of December 2004. b Limit depends on flow in the Rio Grande River.

for process wastewaters from iron and steel manufacturing specify maximum daily and monthly average loadings of cyanide ranging from about 0.0003 to 0.02 pounds of cyanide per 1000 pounds of product (Table 25.3). As indicated in Table 25.3, pretreatment standards are process-specific. They are developed on the basis of process wastewater characteristics and treatability experience. Pretreatment standards have also been developed for cyanide in process wastewaters from a number of industries, for example, at least 55 industrial sectors in the United States ([7]; see also Chapter 18).

25.2 FATE OF CYANIDE IN MUNICIPAL WASTEWATER TREATMENT Influent wastewater to POTWs typically has less (and sometimes much less) than 0.5 mg/l total cyanide, which is predominantly dissolved and in the form of metal–cyanide complexes [4]. While the amount of cyanide coming into a POTW is usually small, the fate of cyanide in the treatment plant is very much of interest because (a) effluent limits for cyanide can be very low, for example, 1 to 20 ppb and (b) cyanide can be inhibitory to the microorganisms in biological treatment at POTWs. A conventional POTW involves primary sedimentation for large particulate matter removal, aerobic biological treatment for removal of biodegradable dissolved and colloidal organic matter and nitrogen, secondary clarification for separation of biosolids, and disinfection by chlorination or UV irradiation. The different chemical forms of cyanide can undergo various kinds of reactions in these processes, many leading to removal of cyanide from solution, but disinfection by chlorination or UV irradiation, as discussed in the next section, can lead to formation of additional cyanide. Fate processes relevant to cyanide in POTWs are [3]: • Removal by adsorption to inorganic particles or biosolids followed by sedimentation • Removal by aeration processes (via “stripping”) and volatilization of HCN

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TABLE 25.3 USEPA Pretreatment Standards for Cyanide in Process Wastewaters Discharged to POTWs from Iron and Steel Manufacturing Operations Max daily mass loading (lb cyanide/1000 lb product)

Avg monthly mass loadingb (lb cyanide/ 1000 lb product)

Source

Source type

40 CFR a section

By-product cokemaking: iron and steel By-product cokemaking: iron and steel By-product cokemaking: merchant By-product cokemaking: merchant Sintering Sintering Iron blast furnace Iron blast furnace Salt bath descaling, reducing: batch Salt bath descaling, reducing: batch Salt bath descaling, reducing: continuous Salt bath descaling, reducing: continuous

Existing

420.15

0.0172

0.00859

New

420.16

0.0172

0.00859

Existing

420.15

0.0200

0.0100

New

420.16

0.0200

0.0100

Existing New Existing New Existing

420.25 420.26 420.35 420.36 420.85

0.00300 0.00100 0.00175 0.000584 0.00102

0.00150 0.000501 0.000876 0.000292 0.000339

New

420.86

0.00102

0.000339

Existing

420.85

0.00569

0.00190

New

420.86

0.00569

0.00190

a USEPA, Code of Federal Regulations, 40, Part 420-Iron and Steel Manufacturing Point Source Category,

U.S. Environmental Protection Agency, Washington, DC, 2002. b Average of daily values for 30 consecutive days.

• Chemical conversion into other species • Aerobic biodegradation. In this section, the fate of cyanide species in each of the major POTW unit processes in a conventional municipal wastewater treatment plant is considered. The reactions of cyanide species upon oxidation in the disinfection process are discussed separately in the next section. In primary sedimentation, in which large particles are removed by gravity settling, opportunity for removal is mainly through interaction of the dissolved cyanide species with the particles. Since the wastewater streams from various sources will often not be blended completely until the wet well at the POTW, the primary sedimentation tank will generally provide the first opportunity for controlled, sustained solid-water contact. Cyanide species may adsorb onto particles in this process. The wastewater particles settled in primary sedimentation are typically 60 to 80% organic, with the remainder inorganic, a large fraction of which is sand and small gravel [8]. Free cyanide and metal–cyanide species can adsorb to solids in wastewater treatment, with the extent of adsorption strongly dependent on the type of solid and solution conditions. Metal–cyanide species are known to adsorb to reactive oxidic minerals such as iron and aluminum oxides [9,10], but are weakly or not at all adsorbed to sand [11]. The adsorption is pH-dependent, with adsorption

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of the anionic metal–cyanide species generally greater at lower pH values and decreasing as the pH is increased. Metal cyanides are also known to adsorb, with pH dependence, to organic matter, including soil organic matter [12] and granular activated carbon [13,14]. In contrast, free cyanide adsorbs very weakly or not at all on oxidic minerals [10,11], but can adsorb to organic matter, including biosolids [12,15,16]. Detailed discussion of the adsorption properties of cyanide species on different types of solids is provided in Chapter 5. Available data indicate limited to no removal of cyanide species in primary sedimentation [3]. There may be some adsorptive removal to inorganic and organic particles, but the extent of removal appears to be small. Volatilization can contribute to cyanide removal in primary treatment as well as adsorption on particles. Influent concentrations of cyanide are often too low to enable accurate assessments of the extent of cyanide removal in primary sedimentation in operating plants. USEPA pretreatment program guidance indicates approximately 27% removal of cyanide in primary treatment [17]. This estimate represents the median percent-removal value measured at 12 POTWs during a 40-POTW USEPA study [18] of removal efficiencies of priority pollutants through primary treatment (i.e., from POTW influent through primary effluent). Only POTWs with average influent concentrations exceeding three times the detection limit of each pollutant were considered for the study. In activated sludge units where aerobic biological treatment takes place, there are several processes that can result in removal of cyanide, including adsorption onto biomass, air stripping, and biodegradation. Generally, removal efficiencies are high and exceed 70% of the incoming cyanide [3]. In a study of cyanide removal in activated sludge units at a plant in California, removals ranged from 60 to 95%, with higher percent removal associated with the higher influent cyanide concentrations as might be expected. The estimated mean percent removal of cyanide by the activated sludge secondary process was 89% [19]. Additionally, cyanide removal efficiencies through activated sludge treatment and other biological POTW processes have been reported by the USEPA as follows [17]: • Activated sludge treatment: ranging from 3 to 99%, with a median removal of 69% • Trickling filter treatment: ranging from 7 to 88%, with a median removal of 59% Air stripping of cyanide through volatilization of HCN will be most pronounced in cases where the water is being agitated and aerated, for example, through surface aerators or bubble diffusers. In experimental bioreactor units with free cyanide in the influent, air stripping was responsible for a large fraction (up to 80%) of the total cyanide removed from the aqueous phase [20]. Microbial degradation also accounted for a substantial (up to 50%) amount of cyanide removal [20]. Removal of cyanide by adsorption on biosolids was small but measurable [16,20]. The extracellular composition of the bacterial cells determines the extent of cyanide adsorption. In similar laboratory tests with properly acclimated biomass, 95 to 99% free cyanide degradation was observed [21,22]. Air stripping is likely to be the predominant removal mechanism in biotreatment systems with unacclimated microbial populations. Free cyanide (including that readily releasable from weak metal–cyanide complexes captured in WAD cyanide analysis) can be metabolized by acclimated aerobic bacteria, but free cyanide can be toxic to unacclimated bacteria. In batch experiments with mixed bacterial populations and glucose as the primary substrate, Zintgraff et al.[23] demonstrated that cyanide is highly toxic to unacclimated cultures. Incremental additions of 0.65 mg/l of free cyanide (as CN) extended the lag phase of unacclimated bacteria by 10 h, and above 6.5 mg/l the growth rate decreased. Free cyanide concentrations as low as 2.6 µg/l (as CN) produced a noticeable effect on the duration of the lag phase of growth in unacclimated cultures. USEPA has reported cyanide inhibition to carbonaceous matter removal at POTWs at cyanide concentrations of 0.1 to 5 mg/l. On the other hand, in experiments with mixed populations living on glucose as the carbon source but acclimated to cyanide, metabolism of cyanide at concentrations up to 10 mg/l proceeded without difficulty [20]. Acclimated heterotrophs have been observed to tolerate concentrations as high as 20 mg/l [21,22], as discussed in Chapter 23.

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Nitrification systems are particularly susceptible to toxic inhibition by cyanide [3]. Depending on the extent of acclimation, nitrification systems can tolerate 0.3 to 2.0 mg/l cyanide, and sometimes more. Wong-Chong and Caruso [22] and Wong-Chong [21] reported inhibition to nitrification in acclimated biomass at 0.5 mg/l free cyanide. A useful guide is that free cyanide concentrations greater than 0.2 mg/l should be avoided for stable operation of biological nitrification processes [24]. The occurrence of adsorption of free cyanide by biomass in aerobic biological treatment means that there is some removal of cyanide from the system in secondary clarification where biosolids are separated from the treated wastewater. Secondary clarifier sludge has not been a major repository of cyanide, but nevertheless the sludge does contain some cyanide and contributes to the overall plant mass balance for cyanide. Kelada [25] measured total and WAD cyanide on biosolid sludges (whole suspension analyzed) from four POTWs. Total cyanide ranged from 491 to 3790 µg/l, while the WAD cyanide was 57 to 436 µg/l. Thus, approximately 90% of the cyanide in the biosolids was in the form of metal–cyanide complexes. Concentrations of total cyanide in effluents from conventional wastewater treatment plants are generally in the range of 0.005 to 0.050 mg/l total cyanide as CN [3,4]. Because of the high degree of removal of cyanide in activated sludge units, concentrations of cyanide in influent to the disinfection process are typically very low to nondetectable. However, at many plants cyanide concentrations in the effluent from the disinfection stage is higher than in the influent. This points to the possibility of cyanide generation in the chlorination (or other oxidation process) used for disinfection, a subject addressed separately below.

25.3 GENERATION OF CYANIDE IN MUNICIPAL WASTEWATER TREATMENT While influent concentrations of total cyanide at POTWs typically very low, cyanide has been periodically detected at elevated concentrations in POTW final effluent, often at concentrations that exceed the influent cyanide concentrations. This has been observed at treatment plants across the United States [4,26,27]. The phenomenon is of concern because water-quality-criteria-driven discharge limits for cyanide are typically very low, often in the range of 1 to 20 µg/l (see Chapter 18). False positives due to analytical interferences can sometimes explain the apparently anomalous detections of cyanide in plant effluent, but careful studies have demonstrated that there are multiple means by which cyanide can be generated in small quantities during the treatment process [4,28, 29], and POTW and commercial laboratories typically employ steps during sample analysis to address known potential interferents (e.g., sulfide). Investigations on the sources of cyanide formation in several POTWs have pointed to two main sources: (1) internal recycle streams from waste biosolids incineration scrubber waters and, (2) the disinfection process, which most commonly is chlorination [4]. Investigations of the nature of these sources have been conducted.

25.3.1 IN-PROCESS CYANIDE GENERATION A comprehensive cyanide monitoring effort conducted in 2000 to 2001 at an East Bay Municipal Utility District (EBMUD) POTW near San Francisco provided clear evidence of cyanide generation in the chlorination process [28–30]. EBMUD sampled weekly at three locations: prechlorination, postchlorination, and postdechlorination for total, WAD, and free cyanide as well as some 35 other analytes. Figure 25.1 shows results of WAD analyses from a sampling sequence conducted on September 24 and 25, 2000. Among the four sampling events, in three cases the chlorinated and dechlorinated effluent WAD cyanide concentrations exceeded those in the unchlorinated effluent [5].

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12 Prechlorination Postchlorination

10

WAD cyanide, mg/l

Postdechlorination 8

6 4 2

AM 9/

25

/0

0

7:

00

AM 0 9/

25

/0 9/

24

/0

0

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00 7:

00

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PM 00

1: 0 /0

24 9/

9/

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/0

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:0

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FIGURE 25.1 Weak acid dissociable (WAD) cyanide monitoring data from a sampling sequence conducted at three locations of EBMUD POTW, September 24–25, 2000. (Source: Kavanaugh, M.C. et al., Report 98-HHE-5, Water Environment Research Foundation, 2003. With permission.)

Zheng et al.[28–30] investigated several chemical reactions (or reaction groups) that may lead to the formation of cyanide in disinfection processes. Among all mechanisms considered, the chlorination of thiocyanate seemed to be the most important for the formation of cyanide in wastewater treatment processes, especially in chlorination/dechlorination, as observed both in laboratory studies and in POTW monitoring. Laboratory experiments have demonstrated that free cyanide can be formed from thiocyanate (SCN− ) upon chlorination where the chlorine dose is not sufficient to completely destroy the SCN− [28]. − + − SCN− + 4Cl2 + 5OH− → SO2− 4 + CNO + 5H + 8Cl

(25.1)

Free cyanide is produced as an intermediate in the reaction shown in Equation (25.1), and subsequently converted to cyanate (CNO− ) if sufficient chlorine is present. Because of the kinetics and the dependence of the reaction progress on the amount of chlorine present, small but measurable concentrations of free cyanide can be present upon chlorination of thiocyanate solutions. This has been observed at the Deer Island municipal wastewater treatment plant of the Massachusetts Water Resources Authority (MWRA), where SCN− and total cyanide concentrations have been monitored simultaneously as shown in Table 25.4. Over a 3-month monitoring period in 2001, total cyanide was consistently nondetectable in the chlorination influent, and consistently detectable at concentrations up to 14.0 µg/l in the postchlorination and postdechlorination samples. At the same time, SCN− present in the prechlorination water at concentrations ranging from 13.7 to 21.8 µg/l was reduced to less than 8.6 µg/l in the postchlorination samples. Zheng et al. [28] also found that thiocyanate can be broken down to yield cyanide by UV irradiation. There is also a possibility that ozonation can convert thiocyanate to cyanide under some conditions [31,32]. The potential formation of cyanide and cyanogen chloride (CNCl) from the chloramination of organic compounds in POTW secondary effluent has been addressed [33]. Chloramines can form

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TABLE 25.4 Summary of Weekly Monitoring Data of Thiocyanate and Total Cyanide and Effects of Dechlorinating Reagents on Cyanide Concentrations in Three Locations at Deer Island Treatment Plant, MWRA, April–June, 2001 (units: µg/l) Postchlorinationa Prechlorination

Postdechlorinationb

CN

CN

Date

CN

SCN

AsA

Na2 S2 O3

SCN

AsA

Na2 S2 O3

SCN

4/19/2001 4/26/2001 5/3/2001 5/10/2001 5/17/2001 5/24/2001 5/31/2001 6/7/2001

NDb

15.2 20.0 17.7 21.8 15.7 17.2 16.7 13.7

ND 7.33 9.67 13.2 7.84 14.0 8.18 11.5

ND 6.13 6.10 8.75 ND 8.03 ND 7.78

ND 5.60 7.20 8.00 7.14 8.63 7.47 7.43

ND 5.35 5.24 7.62 5.15 6.72 6.73 ND

ND 5.44 5.72 7.9 ND 9.09 7.51 ND

ND ND ND 3.7 ND ND ND 5.14

ND ND ND ND ND ND ND

a Postchlorination and postdechlorination samples for CN analysis were dechlorinated with two reagents: ascorbic acid (AsA) and sodium thiosulfate (Na2 S2 O3 ), prior to CN analysis. b ND = not detected. Source: Zheng, A., Dzombak, D.A., and Luthy, R.G., Water Environ. Res., 76, 205, 2004. With permission.

when chlorine in limited dosage reacts with residual ammonia in the treated wastewater. Studies with synthetic solutions demonstrated that CNCl can be formed in significant concentrations by chloramination of various organics such as L-serine, benzene, catechin, and humic acid. Amino acids such as L-serine, which can be exuded from microbes, generally yielded the largest concentrations of CNCl upon chloramination. Studies with POTW secondary effluents demonstrated that CNCl is formed in the chloramination of POTW secondary effluent, which indicates that precursors of CNCl exist in biotreated POTW wastewater [33]. Generally, chlorination of POTW secondary effluent containing residual ammonia can lead to chloramine formation with subsequent chloramination of organic compounds and the resulting production of CNCl and free cyanide at concentrations from 5 to 25 µg/l. CNCl hydrolyzes rapidly, but free cyanide may remain in POTW final effluent if residual chlorine is low.

25.3.2 CYANIDE IN SCRUBBER WATERS FROM ON-SITE BIOSOLIDS INCINERATION On-site biosolid sludge incineration, a common disposal practice for POTWs with limited access to other sludge disposal options such as landfilling, can result in an internal cyanide loading mechanism at POTWs. Specifically, incinerator flue gas scrubber water can contain relatively high (up to 20 mg/l) concentrations of cyanide, and may represent a significant mass loading of cyanide if returned to the main POTW process stream [19,34]. For example, in the Central Contra Costa Sanitary District POTW, Martinez, California, this mechanism has been identified as the only significant source of cyanide to the main wastewater process stream and plant effluent [19,35]. Most of the cyanide in scrubber water typically is free cyanide [34]. Cyanide is likely formed in the multihearth sludge incineration furnaces as a result of localized pockets of reducing conditions at high temperature,

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especially near the inlet stages. Under these conditions, the nitrogen in the sludge (about 5 to 6% by weight) can react with organic matter carbon to form HCN. This HCN is carried off in the flue gas and is removed by the flue gas scrubber water. Cyanide yield at the Cranston, Rhode Island POTW sludge incinerator is about 2 g cyanide per kilogram of dry sludge incinerated [34]. At this plant, the sludge incinerator flue gas is scrubbed with secondary effluent with a continuous flow to the scrubber of about 0.9 mgd. The scrubber water is returned to the plant headworks and merged with influent to the facility. In 2004, the plant’s NPDES discharge limit for total cyanide was 175 µg/l, for discharge to the Pawtuxet River, but a reduction in the discharge limit to 20 µg/l was anticipated [34]. The plant effluent already was below the 20 µg/l level for total cyanide, but efforts were underway to improve the biological treatment operation at the facility for more effective removal of the cyanide associated with the flue gas scrubber water [34]. It has been suggested that operating sludge incinerators at either higher temperatures or higher excess oxygen settings may help to increase the degree of combustion and therefore reduce the amount of cyanide imparted to the wet scrubber stream, but only limited data have been collected to test this hypothesis. Moreover, employing this potential solution would require balancing cyanide formation control objectives with other parameters (e.g., opacity of air emissions) that govern incinerator operations.

25.4 MANAGEMENT STRATEGIES FOR LOW-LEVEL CYANIDE IN DISCHARGES FROM POTWS Numerous wastewater plants across the United States have detected cyanide in chlorinated effluents at levels exceeding those in influent waters, and in some cases, exceeding National Pollutant Discharge Elimination System (NPDES) discharge limits. As discussed in Section 25.3.1, a number of investigations at POTWs have shown that cyanide may be formed during wastewater chlorination or that analytical interferences in monitoring samples can lead to cyanide false positives if not properly evaluated and addressed in sample handling and analysis. The formation of cyanide in chlorination processes is paradoxical, since alkaline chlorination is the leading treatment approach for cyanidebearing wastewaters. The occurrence of this phenomenon is related to a number of factors, including the presence of particular precursor compounds. Interestingly, its emergence appears to coincide with reduction in chlorine doses in order to meet other objectives such as reduced trihalomethane production. The presence of cyanide in raw and treated wastewaters poses numerous challenges to both POTWs and industrial dischargers to POTWs. This is especially the case for: 1. Coastal discharges of wastewater with total cyanide above 1 µg/l, the U.S. water quality criterion for marine environments. 2. Discharges of wastewater with total cyanide in the range of 1 to 20 µg/l, which may exceed local water quality criteria or discharge permit limits if no dilution credit is allowed. 3. The implication of regulatory developments such as the California Toxic Rule (CTR) and similar legislation, which includes severe fines on any discharger exceeding water quality criteria in any treated wastewater sample collected by the discharger. 4. Industrial facilities discharging wastewater to a POTW where pretreatment limits are based on total cyanide as opposed to a measurement of the toxic cyanide species (i.e., free cyanide or weak acid dissociable (WAD) cyanide). Low effluent discharge limits (1 to 20 µg/l) for cyanide are often just above (or even below) the method detection limit (MDL) for the specified type of cyanide analysis. This narrow difference between the MDL and the effluent limit poses a continuing problem for permit compliance because of the potential for false positives due to analytical artifacts and because of the inherently

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larger coefficient of variation associated with results near the detection/reporting limits. There are several strategies that can be pursued by municipal and industrial dischargers to maintain compliance with cyanide discharge standards. Each wastewater discharge case involving cyanide species will have site-specific factors that must be taken into account in developing an appropriate strategy for maintaining compliance. Alternative cyanide compliance strategies for dischargers [5,36] include the following: 1. Reduction of cyanide precursors to sanitary sewers from industrial or commercial dischargers. 2. Obtaining a site-specific discharge permit. 3. Requesting a change in the point of monitoring where compliance would be evaluated. 4. Modification of the activities within a wastewater plant causing production of cyanide, for example, the disinfection process. 5. Requesting a change in the permit requirements with specific attention to alternative analytical methods measuring the concentration of the toxic forms of cyanide.

25.4.1 REDUCTION OF CYANIDE OR CYANIDE PRECURSORS A municipal agency responsible for operating a POTW could undertake a survey of current industrial dischargers and, through characterization of current and future waste streams, determine the magnitude of cyanide species and cyanide precursors such as thiocyanate [4,28] being discharged to the POTW. Pretreatment standards and pretreatment limits could be established to reflect the desire to reduce cyanide or cyanide precursor input to the wastewater treatment plant. Improved monitoring and enforcement may be required if the quantity of cyanide or cyanide precursors cannot be reduced. For example, as noted in Section 25.3.2, some municipal treatment plants generate cyanide internally through practices such as recycling of sludge incinerator off-gas scrubber wash water. In such a case, optimizing the operation of the incinerator may reduce the quantity of cyanide produced in the incinerator off-gas, but may or may not be practical, given furnace conditions, operational flexibility, and other regulations (e.g., air emissions) governing incinerator operations. Further, low-level cyanide production may be an inherent factor of incinerator operation.

25.4.2 REQUEST A SITE-SPECIFIC DISCHARGE PERMIT The 1985 USEPA guidelines [37] for the derivation of numeric aquatic site-specific water quality criteria acknowledged that “… the national criteria, because they are to protect the integrity of all water bodies, serve as benchmarks and may require adjustments for site-specific applications.” One of the options for dischargers is to challenge the applicability of the national water quality criteria for cyanide to the particular receiving water and discharge situation. One such example is a sitespecific standard approved for dischargers in the Puget Sound region in the state of Washington. Water quality criteria previously established were based on a crab species not indigenous to the area. Alternative crab species were tested and a new site-specific water quality criterion was established [38]. Alternatively, one can request a change in the numerical standard based on mixing (dilution) credit analysis or a change in the analytical technique. For example, the Municipal Water Reclamation District of Greater City of Chicago pursued and received approval for an alternative, site-specific water quality criterion that was consistent with relevant aquatic organisms in the receiving water and the appropriate kinds of cyanide analyses [27]. The most commonly used option in site-specific criteria development uses the USEPA indicator species approach to determine Water Effects Ratios (WER), originally developed for analysis of metal toxicity [39]. As of this writing, toxicity in effluent and receiving waters has not been widely evaluated to ascertain whether cyanide exhibits a WER greater than 1 µg/l.

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25.4.3 CHANGE POINT OF COMPLIANCE It has been demonstrated that following chlorination, dechlorination can sometimes lead to a reduction in the amount of cyanide that is either present or has been measured as an analytical artifact [4]. Some POTWs have pursued this approach (e.g., EBMUD; see Ref. [5]), with the approval of governing regulatory agencies, to resolve their cyanide compliance problem.

25.4.4 IN-PLANT MODIFICATION OF OXIDATION PROCESSES Chlorination and UV irradiation of municipal wastewater for disinfection can lead to the formation of amounts of cyanide in the range of typical water-quality-criteria-driven discharge limits [4,28]. Consequently, reductions in effluent cyanide concentrations may be achievable through modifications to the disinfection process. Determination of effective modifications requires site-specific evaluations. In the case of the chlorine disinfection process, two approaches are available: raise the chlorine dose slightly in order to destroy any free cyanide produced, and implement more effective control of the in-process mixing regime to reduce localized occurrences of low pH in the proximity of chlorine addition. The latter will reduce the potential formation of cyanide through reactions such as nitrosation [30] that are favored by low pH conditions, and also localized occurrences of relatively high concentrations of chlorine and therefore localized regions of potentially accelerated/increased cyanide formation.

25.4.5 ALTERNATIVE ANALYTICAL METHODS FOR COMPLIANCE Another strategy that can be pursued is modification of discharge permits by changing the acceptable analytical method from the total cyanide method to an alternative approved method such as the available cyanide method [40], the free cyanide by microdiffusion method [41], or the WAD cyanide method [42], which provide measurement of the more toxic forms of cyanide. This is especially true if it can be shown that the standard based on total cyanide measurements is too conservative and does not provide a benefit commensurate with the increase in costs associated with treatment to meet a total cyanide standard. Discharge standards are intended to provide sufficient protection to human health and the environment without requiring expenditures that yield only limited additional benefits and limited additional risk reduction.

25.5 SUMMARY AND CONCLUSIONS • Cyanide present in influent to POTWs derives from industrial sources, predominantly from electroplating and metal finishing, chemical production, steel production, and petroleum refining. • Regulations usually require pretreatment of industrial wastewaters bearing cyanide prior to discharge to a POTW. • Influent concentrations of total cyanide at POTWs are generally low, ranging from nondetectable to 500 µg/l. The cyanide is primarily (i.e., 60 to 70%) in the form of metal–cyanide complexes. • The fate of cyanide in municipal wastewater treatment depends on the form and amount of cyanide. Free cyanide can be removed by aerobic biodegradation, air stripping, and adsorption to biosolids in secondary wastewater treatment plants, if not present at concentrations toxic to aerobic bacteria (concentrations above 0.2 mg/l are of concern). Understanding of the fate of metal–cyanide species is less well developed. • Investigations at treatment plants have demonstrated that low concentrations (e.g., 5 to 20 µg/l) of cyanide can be formed in POTW disinfection processes, especially chlorination but including UV irradiation. This is a problem as effluent limits for cyanide are often in

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• •

•

513

the range of 1 to 20 µg/l, which are concentrations at or near the method detection limit for most commercially practiced cyanide analytical procedures. Thiocyanate (SCN− ) is an important cyanide precursor in wastewaters. There are two major pathways through which SCN− can be chemically converted to free cyanide: incomplete oxidative destruction by chlorine (due to relatively low chlorine dosing) and by UV irradiation. Chloramination can occur in POTWs when chlorine reacts with residual ammonia present and could be a potential pathway for cyanide and cyanogen chloride (CNCl) formation observed in POTWs that operate without nitrogen removal. On-site biosolids sludge incineration may result in an internal cyanide loading at POTWs. Specifically, incinerator flue gas scrubber water can contain relatively high (up to 20 mg/l) concentrations of cyanide, and may represent a significant mass loading of free cyanide if returned to the main POTW process stream. Strategies for POTWs to maintain compliance with low effluent discharge limits for cyanide include reduction of cyanide precursors through pretreatment, obtaining a site-specific discharge permit, changing the point of monitoring used for compliance, modification of the disinfection process, and changing the cyanide analytical method used for compliance.

REFERENCES 1. USEPA, Iron and steel manufacturing point source category, Code of Federal Regulations, Part 420, Sections 391–412, U.S. Environmental Protection Agency, Washington, DC, 2002. 2. Lordi, D.T., Lue-Hing, C., Whitebloom, S.W., Kelada, N., and Dennison, S., Cyanide problems in municipal wastewater treatment plants, J. Water Pollut. Control Fed., 52, 597, 1980. 3. Wild, S.R., Rudd, T., and Neller, A., Fate and effects of cyanide during wastewater treatment processes, Sci. Total Environ., 156, 93, 1994. 4. Zheng, A., Dzombak, D.A., Luthy, R.G., Kavanaugh, M.C., and Deeb, R.A., The occurrence of cyanide formation in six full-scale publicly owned treatment works, Water Environ. Res., 76, 101, 2004. 5. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Project 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 6. Paschka, M.G., Ghosh, R.S., and Dzombak, D.A., Potential water-quality effects from iron cyanide anticaking agents in road salt, Water Environ. Res., 71, 1235, 1999. 7. USEPA, General pretreatment regulations for existing and new sources of pollution, Code of Federal Regulations, Part 403, Sections 16–18, U.S. Environmental Protection Agency, Washington, DC, 2003. 8. Metcalf and Eddy, Wastewater Engineering: Treatment and Reuse, 4th ed., McGraw Hill, New York, 2003. 9. Bushey, J.T. and Dzombak, D.A., Ferrocyanide adsorption on aluminum oxides, J. Coll. Interface Sci., 272, 46, 2004. 10. Theis, T.L. and West, M.L., Effects of cyanide complexation on the adsorption of trace metals at the surface of goethite, Environ. Technol. Lett., 7, 309, 1986. 11. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 12. Chatwin, T.D., Zhang, J., and Gridley, G.M., Natural mechanisms in soil to mitigate cyanide release, in Proceedings of the Superfund ’88, The 9th National Conference, Hazardous Materials Control Research Institute, Washington, DC, 1988, p. 467. 13. Aksu, Z. and Calik, A., Adsorption of iron(III)–cyanide complex ions to granular activated carbon, J. Environ. Sci. Health, A34, 2087, 1999. 14. Saito, I., The removal of hexacyanoferrate (II) and (III) ions in dilute aqueous solution by activated carbon, Water Res., 18, 319, 1984. 15. Higgins, C.J. and Dzombak, D.A., Free cyanide sorption on freshwater sediment and sediment components, J. Soil Sed. Contam., submitted, 2005.

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16. Raef, S.F., Characklis, W.G., Kessick, M.A., and Ward, C.H., Fate of cyanide and related compounds in aerobic microbial systems. I. Chemical reaction with substrate and physical removal, Water Res., 11, 477, 1977. 17. USEPA, Guidance manual on the development and implementation of local discharge limitations under the pretreatment program, EPA-430/9-87-001, U.S. Environmental Protection Agency, Office of Water, Division of Water Enforcement and Permits, Washington, DC, 1987. 18. USEPA, Fate of priority pollutants in publicly owned treatment works, Volume I, EPA-440/01-82-303, U.S. Environmental Protection Agency, Washington, DC, 1982. 19. MPI, Technical review of cyanide studies at the CCCSD wastewater facility, report to Central Contra Costa Sanitary District, Martinez, CA, Malcolm Pirnie, Inc., Emeryville, CA, 1999. 20. Raef, S.F., Characklis, W.G., Kessick, M.A., and Ward, C.H., Fate of cyanide and related compounds in aerobic microbial systems. II. Microbial degradation, Water Res., 11, 485, 1977. 21. Wong-Chong, G.M., Biological degradation of cyanide in complex industrial wastewaters, in Proceedings of the International Symposium of Biohydrometallurgy, BIOMINET, CANMET Mining and Mineral Sciences Laboratories, Natural Resources Canada, Jackson Hole, WY, 1989. 22. Wong-Chong, G.M. and Caruso, S.C., Biological treatment of by-product coke plant wastewater for the control of BAT parameters, in Proceedings of the Symposium on Iron and Steel Pollution Abatement Technology, EPA-600/9-82-021, U.S. Environmental Protection Agency and American Iron and Steel Institute, Washington, DC, 1982. 23. Zintgraff, G.D., Ward, C.H., and Busch, A.W., Cyanide inhibition of mixed microbial populations, Dev. Ind. Microbiol., 10, 253, 1969. 24. Neufeld, R., Greenfield, J., and Reider, B., Temperature, cyanide and phenolic nitrification inhibition, Water Res., 20, 633, 1986. 25. Kelada, N., Automated direct measurements of total cyanide species and thiocyanate, and their distribution in wastewater and sludge, J. Water Pollut. Control Fed., 61, 350, 1989. 26. Delaney, M.F., Zilitinkevitch, L., McSweeney, N.E., and Epelman, P.E., Cyanide formation from chlorinated POTW effluent, in Proceedings of the Water Environment Federation Environmental Laboratories Conference, Water Environment Federation, Alexandria, VA, 1997, p. 6. 27. Sawyer, B., Zenz, D.R., Lue-Hing, C., Lordi, D.T., and Hill, R., Realistic limits for water toxics: Greater Chicago demonstrates need for site-specific standards, Water Environ. Technol., 10(6), 57, 1998. 28. Zheng, A., Dzombak, D.A., and Luthy, R.G., Effects of thiocyanate on the formation of free cyanide during chlorination and UV disinfection of POTW secondary effluent, Water Environ. Res., 76, 205, 2004. 29. Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Evaluation and testing of analytical methods for cyanide species in municipal and industrial contaminated waters, Environ. Sci. Technol., 37, 107, 2003. 30. Zheng, A., Dzombak, D.A., and Luthy, R.G., Effects of nitrosation on the formation of cyanide in POTW secondary effluent, Water Environ. Res., 76, 197, 2004. 31. Jensen, J.N. and Tuan, Y.J., Chemical oxidation of thiocyanate ion by ozone, Ozone Sci. Eng., 15, 343, 1993. 32. Soto, H., Nava, F., Leal, J., and Jara, J., Regeneration of cyanide by ozone oxidation of thiocyanate in cyanidation tailings, Miner. Eng., 80, 273, 1995. 33. Zheng, A., Dzombak, D.A., and Luthy, R.G., Formation of free cyanide and cyanogen chloride from chlorination of POTW secondary effluent: laboratory study with model compounds, Water Environ. Res., 76, 113, 2004. 34. Bratina, C., Cranston, Rhode Island wastewater treatment plant, personal communication, 2004. 35. MPI, Cyanide test protocol, stage I summary technical memorandum report for Central Contra Costa Sanitary District, Martinez, CA, Malcolm Pirnie, Inc., Emeryville, CA, 2000. 36. Deeb, R.A., Dzombak, D.A., Theis, T.L., Ellgas, W., and Kavanaugh, M.C., The cyanide challenge, Water Environ. Technol., 15(2), 35, 2003. 37. Stephan, C.E., Mount, D.I., Hansen, D.J., Gentile, J.H., Chapman, G.A., and Brungs, W.A., Guidelines for deriving numerical national water quality criteria for the protection of aquatic organisms and their uses, EPA-600/4-85-014, U.S. Environmental Protection Agency, Office of Research and Development, Duluth, MN, 1985.

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38. Brix, K.V., Cardwell, R.D., Henderson, D.G., and Marsden, A.R., Site-specific marine water-quality criterion for cyanide, Environ. Toxicol. Chem., 19, 2323, 2000. 39. USEPA, Interim guidance on determination and use of water-effect ratios for metals, EPA-823-B-94-001, U.S. Environmental Protection Agency, Washington, DC, 1994. 40. USEPA; Method OIA-1677: Available cyanide by flow injection with ligand exchange; U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 64, 73414, 1999. 41. ASTM, Designation D4282-95, Standard test method for determination of free cyanide in water by microdiffusion, Annual Book of ASTM Standards, Vol. 14.02, ASTM International, West Conshochocken, PA, 1998, p. 468. 42. APHA/AWWA/WEF, Method 4500-CN Cyanide, in Standard Methods for the Examination of Water and Wastewater, Clesceri, L.S., Greenberg, A.E., and Eaton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 43. Luthy, R.G. and Jones, L.D., Biological oxidation of coke plant effluent, J. Environ. Eng. Div., ASCE, 106, 847, 1980. 44. Ganczarczyk, J.J., State-of-the-art in coke plant effluent treatment, CRC Crit. Rev. Environ. Control, 13, 103, 1983. 45. Luthy, R.G., Sable, E.R., and McMichael, F.C., Blast furnace recycle water quality and reactions of lead and zinc, J. Water Pollut. Control Fed., 58, 250, 1986. 46. Thiem, L.T. and Alkhatib, E.A., In situ adaptation of activated sludge by shock loading to enhance treatment of high ammonia content petrochemical wastewater, J. Water Pollut. Control Fed., 60, 1245, 1988. 47. Urban, D., Frisbie, S., and Croce, S., Compliance strategy for cyanides in petroleum refinery wastewater: 1. Source characterization and treatment, Environ. Prog., 16, 171, 1997. 48. Palmer, S.A.K., Breton, M.A., Nunno, T.J., Sullivan, D.M., and Surprenant, N.F., Metal/Cyanide Containing Wastes: Treatment Technologies, Corp., N.D., Noyes Data Corp., Park Ridge, NJ, 1988. 49. USEPA, Iron and steel manufacturing point source category, Code of Federal Regulations, 40, Part 420, Sections 391–412, U.S. Environmental Protection Agency, Washington, DC, 2002.

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of Cyanide in 26 Management Industrial Process Wastewaters George M. Wong-Chong, David V. Nakles, and David A. Dzombak CONTENTS 26.1

Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.1 Management Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.2 Regulatory Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.3 Wastewater Characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.4 In-Plant Controls . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.5 Treatment Alternatives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.1.6 Cost Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2 Examples of Cyanide Management in Industrial Wastewaters. . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.1 Hydrometallurgical Gold Mining. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.1.1 Mining Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.1.2 Regulatory Discharge Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.1.3 Wastewater Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.2 Electroplating and Metal Finishing. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.2.1 Process Description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.2.2 Regulatory Discharge Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.2.3 Wastewater Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.3 Blast Furnace Reduction of Metal Ores . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.3.1 Process Description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.3.2 Regulatory Discharge Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.3.3 Wastewater Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.4 Coal Coking . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.4.1 Process Description . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.4.2 Regulatory Discharge Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.4.3 Wastewater Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.5 Petroleum Refining. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.5.1 Refining Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.5.2 Regulatory Discharge Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.2.5.3 Wastewater Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26.3 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

518 518 518 520 521 522 523 523 523 523 524 526 535 535 537 540 541 541 545 545 555 555 555 556 563 563 564 564 566 567

Water and wastewater management are integral components of industrial operations and have a significant impact on the ability of a plant to reliably produce cost-competitive, quality products. Water is used in a number of process operations such as direct and indirect cooling, heating (steam), chemical dissolution, chemical reactions, and general housecleaning. In addition, water may be produced as part of the process operations (e.g., waste ammonia liquor in coking of coal). The net 517

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result of these water uses and production is wastewater, which must be managed in an environmentally acceptable and cost-effective manner. This chapter presents a generic approach for managing cyanide-impacted wastewaters from a manufacturing operation and specific examples of wastewater management in a number of industrial operations where cyanide discharge is one of the primary management issues, that is, gold mining, electroplating and metal finishing, coke production, blast furnace reduction of metal ores, and petroleum refining. Approaches that have evolved for managing cyanide in process wastewaters associated with these industries are described.

26.1 MANAGEMENT STRATEGY Industrial wastewater management in the United States must address compliance with discharge limits, minimizing overall process economics, limited availability of water, and intangibles such as public perception. Figure 26.1 presents a generic strategy for managing the wastewaters from a manufacturing complex that addresses these constraints [1]. As shown, the primary components of this strategy include overall management, waste characterization, in-plant controls, treatment alternatives, and cost and compliance with effluent discharge limits. Regulatory requirements are obviously very important as they dictate input to four of these components: management, waste characterization, treatment alternatives, and most of all effluent discharge quality. For example, specific discharge regulations dictate the levels of controls and extent of treatment required prior to release to the environment (e.g., atmosphere and receiving waters). These regulations also dictate the minimum level of monitoring required to document compliance. The remainder of this section discusses each of the components of this generic management strategy.

26.1.1 MANAGEMENT APPROACH In the past, wastewater management tended to be disconnected from industrial operations. Aqueous waste materials were sent to wastewater treatment, generally without additional thought or timely notification to the wastewater treatment operators. The net results were frequent treatment system upsets and discharge violations. With a more environmentally aware public and the constant business pressure to reduce environmental compliance costs, managers of manufacturing operations are now forced to consider wastewater management as an integral part of the manufacturing operation. To do this effectively, managers of industrial operations had to become more proactive, providing resources and training to execute environmental management plans, and developing a culture that encourages all plant personnel to participate actively in waste management programs. It is increasingly being recognized that effective environmental management plans and programs for industrial process water must include elements such as recycle and reuse, waste minimization, and pollution prevention before end-of-pipe treatment is considered. There are both economic and indirect benefits, such as worker safety and community reputation enhancement. Applying these concepts and integrating them into the corporate and plant operating philosophy can yield both immediate and long-term benefits. For example, Star Plating of New Bedford, Massachusetts, developed a simulated brass coating process to replace its cyanide-based operation and consolidated its other cyanide lines into a single nickel/copper/brass plating line. These operational changes resulted in 30% reduction in cyanide use with concomitant reduction in effluent discharges and near elimination of their cyanide sludge waste stream [2]. These reductions directly translated to cost savings as well as permanent reduction in discharge and disposal liabilities.

26.1.2 REGULATORY REQUIREMENTS Every manufacturing operation in the United States that discharges wastewater to the nation’s waterways is required to comply with some form of discharge limits. These limits are stipulated in

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FIGURE 26.1 Generic industrial wastewater management strategy.

PLANT BOUNDARY

Plant Modifications Capital, Operating and Maintenance POTW Treatment Surcharge Off-Site Treatment Charge Privatization / Outsourcing

Cost

Waste Minimization Integrated (Air, Water, Solid Environmental Control) Recycle/Reuse Process Modifications Water Conservation & Recycling

In-Plant Controls

POTW Receiving Stream

Discharge

Treatment for Recycle/Reuse Discharge to POTW Treatment for Direct Discharge Off-Site Treatment

Treatment Alternatives

Sources Flow Rate, Volume Chemical Composition Physical Characteristics Variations Raw Material Audit Plant Water Usage

Waste Characterization

Upper Management All Plant Personnel

Management

National Pollutant Discharge Elimination System Permit Pretreatment Permit

Regulatory Compliance Monitoring

Clean Water Act Clean Air Act Resource Conservation & Recovery Act Pollution Prevention Act Local Ordinances

Regulatory Requirement

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a discharge permit issued either by a state, as a National Pollution Discharge Elimination System (NPDES) permit for discharge to a receiving water body, or by a local municipality, as part of an ordinance permit for discharge to a Publicly Owned Treatment Works (POTW). These permits limit the allowable levels of contaminants that can be discharged and dictate the monitoring requirements to assure compliance with these limits. Local ordinance limits, guided by federal pretreatment standards, are established to ensure that POTWs are able to comply with their NPDES permits. Generally, NPDES discharge limits are formulated to maintain the quality of the nation’s waterways and are based on United States Environmental Protection Agency (USEPA) industry categorical standards (e.g., Best Available Technology [BAT] effluent limits), and receiving water quality requirements as discussed in Chapters 14 and 18. Similarly, local ordinance limits are also dictated by USEPA categorical pretreatment standards and the water quality criteria governing the receiving stream to which the POTW discharges. Specific categorical limits (industry and pretreatment) are contained in the Code of Federal Regulations (40 CFR Parts 425 to 699). More details on cyanide regulations are presented in Chapter 18 and some specific limits are presented in later sections of this chapter.

26.1.3 WASTEWATER CHARACTERIZATION Critical to any wastewater management plan is a detailed characterization of all wastewater streams. This characterization process, which is often an iterative one, includes identification of sources and locations, analysis of chemical composition, determination of flow rates, and analysis of variability in flow rate and composition. A detailed wastewater characterization requires a plant survey of discharges from each production operation within the overall manufacturing complex. Samples must be collected and analyzed for pertinent constituent parameters, and flow measurements must be made to establish volumetric rates and variability. A survey generally consists of daily measurements over a period of time of 3 to 4 months, or greater, if the production units operate on a campaign basis. Typically, characterization efforts are greater for batch manufacturing operations where the composition of the wastewater can change significantly during the course of a single production campaign and from one production campaign to another. Production campaigns tend to change with market demands for the products, and efforts must be made to project and predict these changes from past production records and market projections. Figure 26.2 shows an example of the degree of variation in chemical oxygen demand (COD) concentration that can occur in the wastewater of a manufacturing plant with time. Design of a treatment system for this wastewater must account for these concentration variations as well as variations in flow rate. The type of manufacturing operation employed will dictate the parameters of concern and the applicable USEPA categorical standards as discussed in Chapter 18. In determining the chemical composition of the wastewater, defensible analytical procedures such as those contained in Standard Methods for the Examination of Water and Wastewaters [3] should be used. In the case of cyanide compounds, it is essential to determine the specific species of the compound (i.e., free, weak acid dissociable [WAD], available, or total cyanide) that is present. Procedures for these measurements are discussed in Chapters 7 and 8. This information on speciation will dictate the type and extent of treatment that is required to achieve regulatory compliance. Water used in housekeeping operations can greatly impact the variability in composition of wastewater, especially in plants where it is an acceptable practice simply to hose-down spilled materials. This type of practice can result in short duration, high concentration spikes which can have adverse effects on a treatment system operation and performance if they are not accounted for during waste characterization. Ideally, the plant survey results will provide a basis for material balances (e.g., water quantity and constituents) around individual production units and the overall facility. However, in older manufacturing facilities, where wastewaters may be discharged through inaccessible, common conveyance

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4,500 4,000 3,500 COD, mg/L

3,000 2,500 2,000 1,500 1,000 500 0

0

10

20

30

40

50

60

70

80

90

100

110

Time, Days

FIGURE 26.2 Example of variations in contaminant concentrations in an industrial wastewater: COD concentration vs. time.

pipes, the survey may yield less than ideal results. The survey data and observations will also provide the basis for the identification of candidate streams for possible product recovery and general recycle and reuse in other sections of the facility, as well as improvements in operating and housekeeping practices. Following the implementation of process modifications and recovery, recycle and reuse options, the characterization process needs to be repeated to evaluate the effectiveness of the modifications and develop a basis for the design of end-of-pipe treatment, should it still be necessary.

26.1.4 IN-PLANT CONTROLS In-plant modification of operation and equipment can be used to recover and minimize waste production through pollution prevention or waste recovery, and to reduce water usage through water recycle and reuse. The results of the plant survey will provide information relative to general water usage and possible production losses that contribute contaminant loads to wastewater streams. These data can be reviewed to determine opportunities for in-process means of eliminating or reducing the discharges by modifying process operations. Opportunities can include: • Changes in process chemicals (e.g., elimination of the use of cyanide as a surface cleaning agent in electroplating or switching to a noncyanide-based plating operation; use of noncyanide lixiviants for gold recovery). • Reduction in operations of water usage through recycle and recirculation (e.g., reuse of tailings water in heap leaching and counter-current cascade; use of rinse water in electroplating). • Installation of product, by-product or feedstock recovery systems (e.g., cyanide recovery from mine tailings and electrochemical recovery of plating metals). • Reuse of treated effluent in manufacturing operations (e.g., cooling tower make-up and landscape irrigation). • Modifications of housekeeping practices to eliminate/minimize water usage and recover spilled product and feedstock. The ultimate goal of applying in-plant controls is to achieve a zero loss of feedstock and products, and zero discharge of wastewater. However, due to the chemical composition of most wastewater streams,

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Neutralization

Filtration

Softening

Membrane Filtration

Ion Exchange

Stripping (Air or Steam)

Chemical Oxidation

Carbon Adsorption

CoagulationPrecipitation

Flotation

Gravity Separation

Biological

Constituents

Reverse Osmosis

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Cyanide (WAD) Cyanide (Complex) Dissolved Inorganics Hardness Heavy Metals

• •

Bulk Trace

Nitrogen( NH3) Oil and Grease Organics

• • • •

Bulk (BOD/COD) Trace/Toxic Refractory Volatile

Phosphorus Suspended Solids pH ( ) Included as a Component of Biological Treatment

FIGURE 26.3 Treatment technology matrix for cyanide in wastewater.

some form of blowdown disposal is almost always needed because of constituent concentration to unacceptable process water quality or excessive production of water. A typical example is the buildup of dissolved salts in cooling tower water [4]. Zero discharge would involve using on-site evaporation or impoundment ponds (e.g., gold mining) or, in the extreme case, multistage evaporation. However, application of the latter approach is expensive and should only be done following a careful economic evaluation.

26.1.5 TREATMENT ALTERNATIVES Based on the quantity and composition of the wastewaters, treatment technologies can be identified to achieve water quality objectives for recycle, reuse, recovery, and discharge. Figure 26.3 presents a matrix of common treatment technologies and the contaminant constituents for which they are effective. Selection of a specific treatment technology will depend on the composition of the wastewater and the quality requirement for its final use or disposition, and will range from the application of simple conventional technologies to complex technology treatment trains. For example, the scrubber effluent water from the wet cleaning of blast furnace off-gas contains high concentrations of suspended solids and traces of ammonia, cyanide, and dissolved metals. Since this scrubber water is used primarily to remove particulate matter entrained in the furnace off-gas, it does not have to be high-quality water. Therefore, it can be recycled following removal of suspended solids by conventional gravity clarification and following pH adjustment to limit the removal of ammonia and cyanide from the off-gas stream; addition of some antiscaling agent will be required to prevent the precipitation of dissolved calcium compounds (e.g., calcium carbonate). However, the direct discharge of this water to a receiving stream would require treatment for ammonia, cyanide, and heavy metals. On the other hand, if the wastewater is highly contaminated, such as coke plant wastewaters (high concentrations of ammonia, cyanide, dissolved inorganic salts, phenol, miscellaneous organics, sulfide, thiocyanate, and heavy metals), water recovery would require a complex treatment train

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involving several technologies, such as: • • • • •

Ammonia-free and fixed leg steam stripper for removal of ammonia, cyanide, and sulfide Biological treatment with denitrification for removal of organics and nitrogen Chemical precipitation for removal of metals, cations, and anions Reverse osmosis with pre-ultrafiltration for removal of suspended and dissolved salt Multi-effect evaporation or other technologies for management of the final brine blowdown.

Ultimately, the final technology selection will be based on plant-specific considerations such as availability of water, ability to achieve the desired treatment goals, overall treatment costs, management of secondary air emissions, and solid/hazardous waste disposal.

26.1.6 COST ANALYSIS Wastewater management alternatives should be evaluated for their cost effectiveness, just like every other component of the manufacturing process. Typical factors that must be considered in any cost analysis include: • Site-specific construction factors (e.g., available power) • Capital cost for new equipment • Operating and maintenance cost for system operation (including labor, chemical and energy usage, residual disposal, and monitoring costs). The cost analysis should also consider the outsourcing of the system operation and the potential costbenefit of changes in manufacturing operations that would result in smaller volumes or improved quality of wastewater. Ultimately, the selection of the wastewater management strategy should be based on achieving the lowest life-cycle cost, highest return on investment, and public acceptance. Patterson [5], Smith and Mudder [6], and Mudder [7] have provided some data on the cost for various cyanide treatment technologies, and cost data are provided for some technologies in Chapters 20 to 23.

26.2 EXAMPLES OF CYANIDE MANAGEMENT IN INDUSTRIAL WASTEWATERS This section presents industrial examples of wastewater management where cyanide is one of the primary chemicals of interest. These examples come from several different industries including gold mining, electroplating and metal finishing, coke production, blast furnace metal ore reduction, and petroleum refining.

26.2.1 HYDROMETALLURGICAL GOLD MINING 26.2.1.1 Mining Process Today’s commercial gold mining and recovery process essentially involves the following operations: • Removal of ore from the mine (underground or open-pit) to a staging area • Ore preparation (e.g., crushing, milling, and flotation) • Aerated leaching of the prepared ore with an alkaline 0.02 to 0.05% solution of sodium cyanide (∼1 to 2.5 kg NaCN/ton of ore) [8] • Recovery of gold and other metals from the cyanide solution by either carbon adsorption or powdered zinc precipitation

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• Partial recycle of cyanide solution, following metal removal, with the remaining portion sent to tailings ponds for solids separation and storage. Figure 26.4 and Figure 26.5 present schematic process diagrams of the hydrometallurgical gold mining operation. These figures show processing operations (e.g., crushing, milling, and flotation) that address the geologic diversity of materials mined, technologies required and available for ore preparation, cyanide leaching (cyanidation process), and the central gold recovery process. Summary details on the technologies for gold mining and processing are available in World Gold [9]. The aerated leaching and gold/metal recovery steps are parts of the cyanidation process, which was first developed in 1877 by J.S. MacArthur, R. Forrest, and W. Forrest [9,10]. In the process, gold is dissolved in an aerated solution of sodium cyanide after which the gold in solution is recovered by zinc precipitation. In many operations, carbon adsorption has replaced zinc precipitation. Cyanidation is extensively used in heap leach mining, where the sodium cyanide solution is sprayed onto an ore pile and leaches the gold while trickling down through the ore. The cyanidation process has made the mining of low-grade ores economically feasible. For example, at the Ruby Hill Project of Homestake Mining in Nevada, ores that contain 0.06 ounces (1.8 g) of gold/ton of ore are being processed [10]. Figure 26.6 shows the quality of ore being mined at various locations in the United States, the process that is used to mine it, and the projected reserves for each ore [11]. This figure shows several mines recovering gold from ores containing less than 1.4 g of gold/metric ton of ore. In a mining operation, wastewaters include acid mine drainage, run-off from ore stockpiles and waste rock stacks, leachate from disposal areas, barren water, tailings impoundment water and leachate, and domestic sewage [6]. • Acid mine drainage: Metals, WAD cyanide, thiocyanate, total suspended solids (TSS), and acidity • Barren water or tailings impoundment water: Metals, WAD cyanide, ammonia, thiocyanate, TSS, alkalinity, and pH • Tailings impoundment leachate: Metals and WAD cyanide • Run-off from ore stockpiles and waste rock stacks: trace metals, TSS, and possibly acidity. Table 26.1 presents chemical composition of barren water and leachates from tailing impoundments at the Lead, SD mine of Homestake Mining. Table 26.2 presents chemical characteristics of tailings waters, slurries, and seepage from several mines including the Homestake Mine, Lead, S.D.; the Martha Hill Gold Mine, Waihi, New Zealand; and the Golden Cross Mine, Australia. The data in these two tables show wide variations in wastewater composition due to the inherent differences in ore mineralogy at the different locations, and the natural heterogeneity of ore at the same mine location. These variations, especially those at the same location, must be considered carefully to ensure proper management of gold mining wastewater. Table 26.1 also presents data on flow rate and temperature. These data are important since temperature impacts treatment reaction rates, with lower temperatures resulting in slower reaction rates that require larger reactors to achieve a specific level of treatment. Similarly, wastewater flow rate also impacts the size of a treatment system. A flow rate ranging from 1 to 6000 gpm naturally leads to challenges when designing a treatment facility, and will require accommodations such as modules to bring online for flows in certain ranges, storage for flow equalization, or others. 26.2.1.2 Regulatory Discharge Requirements Gold mining in the United States is regulated by the Mining Law of 1872. In 1994, the U.S. Congress, in attempting to revise this legislation, was in disagreement on several issues. One of the areas of contention was whether the revised mining law should have provisions for reclamation and

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Smelt to Dore Bullion

Zn Dust Precipitation

Merrill–Crowe Deaeratiton

Cyanide Leach

Regular Heap Leaching

Pretreatment Methods (see Figure 26-5)

Refractory Ores

Crushing

Milling Ores

Filter

Zn Dust Precipitation

Strip Carbon

Electrowinning

Merrill–Crowe Deaeratiton

Cyanide Leach

Nonrefractory Ores

Smelt to Dore Bullion

Pretreatment-In-Pulp / Carbon-In Leach

Grinding

Gravity Separation

FIGURE 26.4 Schematic diagram of gold ore processing. (Source: Data from World Gold, Report SP 24-94, U.S. Department of Interior, Bureau of Mines, Washington, DC, 1994, Figure 2.)

Electrowinning

Strip Carbon

Carbon-In Columns

ROM Dump Heap Leaching

Bio oxidation

Agglomeration

Run-Of-Mine Ore

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Refactory Ores from Grinding (see Figure 26.4) Flotation

Concentrate to Oxidation Methods

Concentrates to Concentrates Off-Site to On-Site Smelter / Refinery Cyanide Leach

Concentrates to Off-Site Cyanide Leach

To Cyanide Leach (see Figure 26.4)

Concentrate Roasting

Oxidation Methods

“Whole Ore” Roasting

Pressure Oxidation (Autoclaves)

Chlorination

Bacterial Oxidation "Biooxidation"

FIGURE 26.5 Pretreatment processes for sulfide ores. (Source: Data from World Gold, Report SP-94, U.S. Department of the Interior, Bureau of Mines, Washington, DC, 1994, Figure 3.)

environmental issues or whether other state or federal laws should prevail [9]. State legislation still governs water quality and management with respect to gold mining. There are no cyanide limitations for gold mining discharges in the U.S. Code of Federal Regulations, Effluent Guidelines and Standards (40CFR Part 440.103). However, this regulation stipulates zero discharge except in cases of excess precipitation run-off, which is allowed to be discharged without a cyanide limit. Similarly, the Australian mining industry faces no national regulations covering the management of cyanide [12]. However, in the United States, gold mining NPDES discharges are regulated state-by-state on the basis of receiving water quality criteria. For example, the Homestake Mine in Lead, South Dakota discharges to Whitewood Creek, which is classified as “cold water marginal fish life propagation.” To secure their permit to discharge, Homestake was required to implement site-specific water quality criteria testing, which included toxicological testing [13]. 26.2.1.3 Wastewater Management Strategy Wastewater management in the gold mining industry has been limited in large measure to impoundments for storage and treatment of the various types of wastewaters listed earlier. These

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0.3 Underground Mining Ore Grade (ounces gold per ton)

0.25

Milling Operations

Jerritt Canyon Homestake

Heap Leaching 0.2 Pinson Battle Mountain

0.15 Alligator Ridge

0.1

Carlin

Borealis Northumberland Golden Sunlight Ortiz

0.05

Smoky Valley

McDonald

Zortman-Landusky

Ridgeway

Montana Tunnels

0 1.E+06

1.E+07

1.E+08

1.E+09

Reserves (tons)

FIGURE 26.6 Comparison of gold reserve quality at various mine operations throughout the United States. (Source: Young, C.A. in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., TMS, p. 97, 2001. With permission.)

impoundments have enabled effective management of the large volumes of wastewater involved in hydrometallurgical gold mining, but they have also been the sources of some large uncontrolled and sometimes catastrophic discharges. Table 26.3 lists examples of such discharge events that have occurred since 1990. These events have been in part due to improper impoundment design, and in many cases these design faults have been compounded by flooding [12,14,15]. Compilations of best practices for treatment of wastewaters in the gold mining are provided in the 1998 Cyanide Monograph [7], the 1991 The Chemistry and Treatment of Cyanidation Wastes [6], and the 2001 Cyanide Social Industrial, and Economic Aspects [11]. These reports cover a number of topics, including: • • • • •

Cyanide management issues Alternative lixiviants Recovery and reuse of cyanide End-of-pipe treatment alternatives for cyanide Spent heap leach pad closure

Best management practices for the cyanidation process are presented, including approaches for environmental responsiveness in a cost-effective manner. Approaches for modifying the gold recovery process to reduce the use of cyanide and for recovery and recycle of cyanide are addressed. The remainder of this section discusses the initial four items in the context of management strategies for gold mining wastewaters. The closure of heap leach pads is discussed in Chapter 27. 26.2.1.3.1 Management issues In spite of all the work by the gold mining industry on managing environmental impacts, the environmental management record of the industry suffers due to the widely publicized cyanide spills and tailings pond failures that have occurred in the past and continue to occur (Table 26.3). Cyanide is the constituent focused upon by the public when the releases occur. The problematic public image of the gold mining industry with respect to environmental management is partly due to poor operating

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TABLE 26.1 Chemical Composition Ranges for Barren Water, Wastewater Decant, and Seepage at Gold Mining Operations Parameter Arsenic Cadmium Chromium Copper Iron Lead Manganese Mercury Nickel Selenium Silver Zinc Total cyanide WAD cyanide Free cyanide Ammonia-N Thiocyanate pH, SU units Hardness (CaCO3 ) Sulfate Temperature (◦ C) Flow (in gpm)

Range of concentrations (mg/l) <0.02–10.0 <0.005–0.02 <0.02–0.1 0.1–400.0 0.50–40.0 <0.01–0.1 0.1–20.0 <0.0001–0.05 0.02–10.0 <0.02–6.0 <0.005–2.0 0.05–100.0 0.5–1000.0 0.5–650.0 <0.01–200.0 <0.1–50.0 <1.0–2000.0 2.0–11.5 units 200–1500 5–20,000 0–35◦ C 1–6000 gpm

Source: Data from Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. With permission.

procedures at some mines [14–16], which have led to prohibitions on the use of the cyanidation process in Montana, the Czech Republic, and Vilas and Oneida countries in Wisconsin [17]. Faced with these threats to the industry, improved implementation of operational and environmental practices for best management is imperative and being addressed by the industry. Environment Australia has recommended the following [12]: • Develop and implement site-specific, best practice of cyanide management plans. These plans should be in place prior to and at the outset of mine operations, should be revised continuously as mine operation progresses, and should address mine closure and rehabilitation following operation. • Provide planners and engineers with specific instructions to design mine operation for safety, environmental responsiveness, and optimum use of cyanide. • Provide resources, training, and possibly incentives for all employees and on-site contractors who handle or are exposed to cyanide. • Establish well-defined responsibilities for each individual involved, with clear chains of command and effective lines of communication.

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TABLE 26.2 Chemical Composition of Gold Mining Process Wastewaters from Specific Mines Mine and source (concentrations, mg/l)

Parameter Arsenic Cadmium Chromium Copper Iron Lead Manganese Mercury Nickel Selenium Silver Zinc Total CN WAD CN Free CN Ammonia–N Thiocyanate Hardness (CaCO3 ) Sulfate pH, SU units Temperature, ◦ C

Golden Cross Mine

Homestake, Lead, SD tailings water

Martha Hill tailings slurry

tailings water

NA NA NA 1.4–6.0 1.5–6.0 NA NA NA 0.2–0.4 NA NA 0.01–1.50 7.0–30.0 4.0–20.0 NA <1.0 110–350 400–500 NA 7.0–9.5 1–21

0.8 <0.01 0.02 4.7 1.3 <0.1 0.01 0.016 NA NA 0.15 0.64 218 213 NA NA 34 307 360 10.4 NA

NA <1.0 <0.02 NA 1.0–5.0 0.001–0.1 0.05–0.1 0.001–0.009 0.01 0.01–0.4 NA 0.5–3.0 150–220 150–220 NA NA 30–40 750 1000 10–11 NA

tailings seepage NA <0.005 0.002–0.02 5.0–10.0 5 0.001 0.03–0.05 0.001–0.005 NA <0.05 NA 2.0–4.0 50 20–30 NA NA NA NA 1000–2000 9.5–10.0 NA

NA: No data available. Sources: Data from Mudder, T., in Cyanide Monograph, Mining Journal Books, Ltd., London, 1998; and Whitlock, J.L. and Mudder, T., in Cyanide Monograph, Mining Journal Books, Ltd., London, 1998.

TABLE 26.3 Recent Cyanide Spills and Tailing Pond Failures Location South Carolina, USA Colorado, USA Guyana Nevada, USA South Dakota, USA Spain Romania Ghana, Wassa West district

Mine Brewer Mine Summitville Mine Omai Mine Gold Quarry Mine Homestake Mine Los Frailes Zinc Mine Aural Mine Goldfields Ltd.

Type of discharge Tailings Tailings, spill Tailings, spill Tailings, spill Tailings, spill Tailings, spill Tailings, spill Tailings, spill

Year 1990 1990 1995 1997 1998 1998 2000 2001

Source: Data from WSN, Group support new bill banning cyanide use in Wisconsin mines, Wisconsin Stewardship Network, http://www.nocrandonmine.com/wsn/mining/cyanidebannews. html, 2004.

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• Institute safety procedures for the transport, storage, containment, use, and disposal of cyanide. • Establish a cyanide management strategy as part of the mine’s environmental management plan. • Integrate the cyanide and water management plans for the mining operation. • Identify, evaluate, and implement appropriate options for reusing, recycling, recovering, and disposing of residual cyanide from plant operations. • Conduct regular cyanide audits and revise cyanide management procedures where appropriate. • Develop and implement occupational and environmental monitoring programs, supported by routine sampling, analysis, and reporting protocol. • Develop a carefully considered site-specific emergency response plan. • Develop an active public education program. Some or all of these practices are in-place at gold mines in operation throughout the world. However, the continuing occurrences of large releases of cyanide-bearing waters from some mines indicate that improved practice of water and wastewater management techniques is needed in the industry. Environment Australia [12] emphasizes that adoption of best practice principles and a proactive approach to cyanide management are likely to derive many benefits, including: (1) Better relationships with public and regulatory agencies; (2) Improved economic and environmental performance; (3) Reduced risks and liabilities; and (4) Easier access to capital and potentially lower insurance costs. 26.2.1.3.2 Alternative lixiviants The cyanide ion (CN− ) binds strongly to gold, making it an effective extraction agent, or lixiviant. In the cyanidation process, CN− binds with elemental gold in the ore according to the following reaction: − 4Au + 8CN− + O2 + 2H2 O → 4Au(CN)− 2 + 4OH

(26.1)

or, 1 1 − Au + 2CN− + O2 + H2 O → Au(CN)− 2 + OH 4 2

(26.2)

These equations indicate that the removal of every gram of gold requires 0.26 g of cyanide ion. In practice, about 0.5 to 1.3 g of cyanide ion is used per ton of ore [8]; mined ores can contain from 1.4 to 5.6 g (0.05 to 0.20 ounces) of gold per ton of ore (see Figure 26.6). Laboratory-scale investigations into alternative lixiviants have been conducted [18–23] and their testing continues at the pilot-plant scale [19,23]. The alternative lixiviants investigated include − − − ammonia (NH+ 4 /NH3 ) [19], bromide (Br /Br2 ) [20,22], chloride (Cl /OCl ) [20,22], iodine (I− /I2 ) [20], catalyzed nitrogen species (NSC/Cl− ) [24], sulfide/polysulfide S−2 /S−2 x 0 [21,24], −2 thiosulfate (S2 O3 ) [18,22,23], and thiourea (CS(NH2 )2 ) [20,22]. Also bioleaching, utilizing Thiobacillus- and Leptospirillum-like bacteria, has been tested at the laboratory and pilot-scale [18]. In one of these bioleaching tests, the reduction of sulfide content by the biological oxidation of pyritic sulfide to sulfate has been tested as a first stage of processing. In a second biological stage, the pyrite-free ore is subjected to treatment, where the sulfate formed in stage one or in acid mine water is biologically reduced back to bisulfide using naturally-occurring bacteria and an added carbon source (e.g., wood or grain alcohol or acetic acid). The biologically generated bisulfide serves as the leaching agent for gold recovery [18]. This biotechnology for gold leaching is still in the experimental stage.

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TABLE 26.4 Concentration Requirements for Alternative Lixiviants for Gold Extraction Leach test concentrations (mg/l) and pH conditions Compiled literature Reactants Cyanide Ammonium thiosulfate Thiourea Sodium bromide/bromine Chloride/hypochlorite Iodide/iodine NSC/chloride Sulfate/polysulfide Ammonium/ammonia

McDonald Ore

Leach concentrations

pH

500 15,000 1250–2940 10,000/1000 100,000/1000 NA NA NA NA

150 11,800 10,000 40,000 35,000 30,000 5000/25 3200 500

11.0 8.0 1.5 2.0 2.5 3.0 1.0 11.0 10.0

Sources: Data from Veto, R.H. and McNulty, T.P., in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 83; and Young, C.A., in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 97.

Tables 26.4 to 26.6 present comparisons of the concentrations of alternative lixiviants required for gold recovery, the gold recovery achieved with these reactants, and the chemical characteristics of the leachates that result from their use. These data show the following: • The chemical requirement for cyanide is significantly less than all other lixiviants except ammonia (500 mg/l vs. 1,000 to 100,000 mg/l). • Gold recovery using cyanide is greater than that produced by other lixiviants (73 vs. 38 to 68%). • The leacheate chemistry for cyanide yields significantly less contaminated leach water than that produced by other lixiviants; leachates of alternative lixiviants are also potentially more toxic than the cyanide leach. Other reported observations supporting the benefits of cyanide processing include: • Implementation of alternative lixiviant processing in heap leaching would require four reusable or “on–off” leach pads as compared to a single leach pad for cyanide [22]. • Because of the high concentrations of alternative lixiviant required for gold leaching, the recirculating leach solution would require continuous treatment to prevent the buildup of dissolved salts, which could retard leaching rate and impair heap porosity and even gold recovery [21,22]. Tests have indicated that a blowdown rate of 10% may be required to maintain adequate processing quality for selected alternative lixiviant [22]. • Gold recovery processes such as carbon-in-leach (CIL), carbon-in-pulp (CIP), and zinc precipitation have not been successful with alternative lixiviants [25]; evaluation of ion exchange continues [26]. • Bioleaching is considered experimental at this time.

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TABLE 26.5 Gold Recovery from Various Lixiviant Solutions Lixiviant

Gold recovery (%)

Cyanide Ammonium thiosulfate Bromide/bromine Chloride/hypochlorite Thiourea

73 38 57 68 57

Source: Data from Veto, R.H. and McNulty, T.P., in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 83.

TABLE 26.6 Leachate Chemistry of Alternative Lixiviant Solutions for Gold Extraction

Parameter (mg/l)

Cyanide system

Thiourea system

Ammonium thiosulfate system

Chlorine system

Bromine system

pH Total dissolved solids (TDS) Sodium Calcium Chloride Sulfate Ammonia, (mg/l as N) Cyanide, total (mg/l as CN) Chlorine, free as Cl2 Bromine, free as Br2 Bromide Thiourea Thiosulfate Aluminum Arsenic Cobalt Copper Iron Manganese Nickel Thallium Zinc

11.0 790 234 64.5 430 10 3.41 280 NA NA NA NA NA 0.290 0.048 0.05 0.29 0.68 <0.005 0.52 0.0004 0.74

1.6 15,100 40 534 1500 11,400 9.31 <0.01 NA NA NA 2000 NA 230 11.0 3.1 2.1 1390 88.50 12.7 0.31 3.1

9.1 30,100 2160 53 10,400 600 5060 <0.01 NA NA NA NA 10,859 <0.2 0.605 0.11 47.3 0.13 0.40 2.77 0.218 0.28

5.9 107,000 36,900 910 66,000 260 <0.05 <0.01 424 NA NA NA NA <3 1.4 <1 0.6 <1 <0.5 16 0.09 33

2.4 15,500 2120 816 5700 3460 1.72 <0.01 NA 1100 8830 NA NA 161 18.1 0.5 0.9 154 32.70 6.6 0.058 4.5

NA: No data available. Source: Data from Veto, R.H. and McNulty, T.P., in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 83.

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Because of the current shortcomings listed above, none of the alternative lixiviants that have been tested to date show economic promise when considering both environmental impacts and economic benefits. As such, cyanide remains the lixiviant of choice [27]. 26.2.1.3.3 Recovery and reuse of cyanide Cyanide recovery was practiced in the mining industry as early as 1930. Mudder and Goldstone [28] present a concise history of the use of air stripping for cyanide recovery. In the early days, recovery was economically driven because of the limited production of sodium cyanide; today, it is driven by both economics and environmental concerns. It is estimated that the cost of using cyanide in gold recovery is about $3.00/ton of ore (cyanide usage of 2.5 kg/ton of ore processed and cyanide cost of $1.20 /kg). It has been estimated that 20 to 30% of these costs can be saved by implementing cyanide recovery [29] in addition to the cost for downstream treatment. Cyanide can be recovered from numerous processing streams in the gold mining operation: barren solution and tailings, tailings slurry, pregnant solution, and mill slurry before cyanidation. Recovery of cyanide from pregnant liquor prior to CIP processing has been shown to increase gold recovery by reducing the competition of compounds, other than the gold complex, for active sites in carbon adsorption; cyanide is a competing compound for adsorption sites [29]. Importantly, though, cyanide oxidation by oxygen from the atmosphere is catalyzed by activated carbon and is a major source of about 50% of the cyanide losses in the gold recovery process [29]. The most common cyanide recovery process is air stripping. This technology has been applied in gold mining in a variety of process configurations (e.g., batch and continuous flow operations, packed columns, and tank reactor). Details of the technology and its process variations (i.e., Mill Crow, AVR, and CYANISORB processes) are discussed in Chapter 21. It must be noted that cyanide recovery by vapor stripping will impact only WAD cyanide and in many cases will not produce water quality that meets cyanide discharge requirements for cyanide. Further, strongly metal-complexed cyanide, SCN− , and the heavy metals (e.g., copper, nickel, and zinc) are not affected by the cyanide recovery process. 26.2.1.3.4 End-of-pipe treatment alternatives The most widely practiced end-of-pipe treatment for gold mining wastewaters has been impoundment ponds utilizing natural processes (e.g., photodissociation, biological decomposition, metal complexation and precipitation, and volatilization) to remove cyanide and heavy metals from the waste. Technically, this treatment approach can be effective, provided that the impoundments are designed and operated properly to meet anticipated flow and chemical loadings as well as severe rainfall events. In some industries, however, treatment provided by impoundments is considered only as partial or pretreatment. The catastrophic failures of some impoundments at active or former gold mining sites has adversely marked this technology as inadequate for management of gold mining wastewater. In addition, protection of wildlife such as waterfowl and deer from these impoundment ponds is a frequently raised issue (Chapters 15 and 17). In the search for alternative treatment options to supplement or replace surface impoundments, a great deal of information on the treatment of cyanide in gold mining wastewaters has been generated and is available [6,11,28,30]. Table 26.7 presents a listing of the various technologies available for treatment of gold mining wastewaters, as well as a summary of application status and example installations where the technology is in use. Details of the various technologies are presented in the documents referenced above and in earlier chapters (Chapters 20 to 23). Under the global heading of treatment technologies, recycle and recovery technologies are included; these technologies are discussed earlier in this chapter and in Chapter 21. The most widely practiced active treatment technology for free cyanide is oxidative destruction with chlorine or hydrogen peroxide, with a preference for chlorination, which provides a wider range of treatment (Chapter 20). Biological treatment (Chapter 23), which has been recognized as the most cost effective in treating other industrial wastewaters, has not gained wide acceptance in the gold mining industry, in part because of technical limitations associated with the presence of

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Biological

Chemical oxidation (1) Chlorination (2) Peroxide (3) SO2 /Air Impoundments Air stripping Carbon adsorption

Ion exchange Membrane separation Solvent extraction

Comm

Technology

Status






Expt’l

















Free CN















WAD CN








Metals

Contaminants

Baker Mine OK Tedi Mine Lynngold Mine All mines Beaconsfield Gold Mine

Homestake Mine

Example installation

TABLE 26.7 Listing of Technologies Practiced and Evaluated for the Treatment of Gold Mining Wastewaters

Partial treatment Recovery technology Not used in wastewater but technology familiar to industry Recovery technology Recovery technology Recovery technology

Incidental removal of SAD and metals

Comments

534 Cyanide in Water and Soil

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strong metal–cyanide complexes. Four applications of biological treatment have been reported at Lead, South Dakota, south central British Columbia, Canada, Green Spring Gold Operation Ely, Nevada [31], and Peak Gold Mine, NSW, Australia [32]. Impoundments used in management of gold mining wastewater, which provide passive treatment for both cyanide and metals, are considered as part of the gold mining operation in that they serve as a recirculation point in the heap leaching process loop. Given the concerns over structural failures and safety to wildlife, an obvious question is whether a more controllable closed-vessel system for process water could be used and be economic when all factors are considered, including: delay costs associated with permitting, issues with impoundment failures and wildlife safety, and impoundment closure. A closed-vessel system might include solids handling and dewatering along with water handling. Some of the cyanide treatment might be lost because of reduced retention time but treatment rates could also be accelerated for certain processes through the better control possible in closed vessels. Impoundments will still be required for general site run-off containment; this water could be used as make-up water and returned to the processing operation. Figure 26.7 presents a conceptual flow diagram of a closed impoundment system. This figure shows front-end processing to remove suspended solids in barren slurry; solids are returned to the mining operation. The partially treated water is stored in covered tanks, possibly in the capacity range of 5 million gallons. This storage allows for eventual recycle to the heap leach operation, and excess water treatment for discharge.

26.2.2 ELECTROPLATING AND METAL FINISHING 26.2.2.1 Process Description Metal finishing by electroplating alters the surfaces of metal products to enhance their mechanical, chemical, electrical, and physical properties, such as corrosion and wear resistance, electrical conductivity and resistance, chemical and tarnish resistance, ability to bond to rubber, reflectivity, and appearance [33]. Table 26.8 presents a list of typical plating metals and their functions. Metal finishing is part of many manufacturing operations, including automobiles, airplanes, kitchen appliances, telecommunication equipment, jewelry, and heavy equipment [2]. There are numerous specialty plating and electroplating operation “shops” that provide contract support to larger manufacturing companies and serve particular market niches. Metal finishing, which is applied to both metallic and nonmetallic surfaces, involves the following operations [33]: • Surface preparation • Mechanical: sandblasting and grinding • Chemical: cleaning, degreasing, pickling, and etching • Thermal, electrical, and sonic • Surface plating and finishing • Chemical: electroless plating • Electrochemical: anodizing and plating • Final polishing • Mechanical: buffing and polishing • Electropolishing Table 26.9 presents an illustration of three different versions of a typical plating cycle; slightly different sets of steps are required depending upon the condition of the material that is to be plated. Note that the actual plating step does not take place until Step 15. Figure 26.8 presents a flow diagram of a typical plating line, in this case for chrome plating of steel. The line consists of electrolytic and chemical cleaning prior to a nickel plate base followed by the final chromium finishing, and involves 11 processing steps.

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Sludge Dewatering

Clarifier

FIGURE 26.7 Conceptual closed impoundment system for gold mine process wastewater.

To Mining Operation

Sludge Cake

Barren Water/Slurry

Hydrocyclone

Polymer

To Leach Pad

Impoundment Tank (s)

To Treatment for Discharge

536 Cyanide in Water and Soil

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TABLE 26.8 Typical Plating Metals and their Functions Property/function Decoration Corrosion resistance Wear, lubricity, hardness Bearings Joining, soldering, brazing, electrical contact resistance, conductivity Barrier coatings, antidiffusion, heat-treat, stop-off Electromagnetic shielding Paint/lacquer base, rubber bonding Manufacturing, electroforming Manufacturing, electronic circuitry Dimensional buildup, salvage of worn parts

Typical plating metals Chromium, copper, nickel, brass, bronze, gold, silver, platinum-group, zinc Nickel, chromium, electroless nickel, zinc, cadmium, copper and copper alloys, gold Chromium, electroless nickel, bronze, nickel, cadmium, metal composites Copper and bronze, silver and silver alloys, lead–tin Nickel, electroless nickel, electroless copper, copper, cadmium, gold, silver, lead–tin, tin, cobalt Nickel, cobalt, iron, copper, bronze, tin–nickel Copper, electroless copper, nickel or electroless nickel, zinc Zinc, tin, chromium, brass Copper, nickel Electroless copper, copper, electroless nickel, nickel, electroless gold, gold Chromium, nickel, electroless nickel, iron

Source: Schlesinger, M., Kirk–Othmer Encyclopedia of Chemical Technology, online edition [June 18, 2004], John Wiley & Sons, New York, copyright 2004. John Wiley and Sons, Inc. Reprinted with permission of John Wiley & Sons, Inc.

Table 26.10 presents a list of the various metals used in electroplating and metal finishing and those metals that require cyanide in their formulations. Many different chemicals are used in the aqueous plating baths employed in electroplating. Cyanide has long been used in plating bath formulations because of its ability to complex with and hold metals in solution, as well as its ability to deposit coatings having the desired appearance and physical properties. Industry interest in “greener” electroplating technologies has spurred the development of a number of alternative, noncyanide formulations. Full commercialization of some of these formulations remains questionable because of the process economics and product quality [34]. As indicated in Table 26.10, cyanide remains commonly used in plating bath formulations.

26.2.2.2 Regulatory Discharge Requirements Electroplating and metal finishing uses a wide range of chemicals for both cleaning the work pieces and the actual plating operation. The chemicals used in these operations can adversely impact human health and the environment, if not properly managed. The metals used in electroplating, such as cadmium, chromium, copper, lead, nickel, silver, and tin, are of environmental concern. In addition, since many plating formulations are cyanide-based, there is obvious concern for cyanide in the discharges. These constituents in electroplating wastes, including spent plating baths and rinse water, are regulated via the effluent discharge limits of the Clean Water Act and the sludge requirements of the Resource Conservation and Recovery Act (RCRA). The Clean Water Act effluent discharge limits (40 CFR Part 413) are written for sources discharging to POTWs, which are divided into two categories: sources discharging less than 38,000 L of treated wastewater/day, and those discharging more than 38,000 l/day. In addition, the regulations are written for the various subcategories of plating operations. Table 26.11 presents a summary of

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TABLE 26.9 Processing Stages in a Typical Electroplating Cycle for Plating Nickel on Low-Carbon Steel Stationa 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19

Heavy soil rust and scale Soak cleanb Rinse Electroclean Rinse Rinse Acid picklec Rinse Rinse Electrocleand Rinse Rinse Acid drip, mild Rinse Rinse Plate nickel Saver-rinse Rinse Rinse Hot rinse and dry

Heavy soil no rust or scale

Light soil no rust

Soak cleanb Rinse Electroclean Rinse Rinse Acid dipc Rinse Rinse

Electroclean Rinse Rinse Acid dipc Rinse Rinse

Plate nickel Saver-rinse Rinse Rinse Hot rinse and dry

Plate nickel Saver-rinse Rinse Rinse Hot rinse and dry

a For high carbon (0.4% C and more), the desmutting steps 9 to 14 are required. b Preclean option maybe required before soak. c For high carbon steel, acid treatment should be as mild as possible. d Alkaline desmut option.

Source: Schlesinger, M., Kirk–Othmer Encyclopedia of Chemical Technology, online edition [June 18, 2004], John Wiley & Sons, New York, Copyright 2004. John Wiely and Sons, Inc. Reprinted with permission of John Wiley & Sons, Inc.

the pretreatment standards for existing sources (PSES) discharge limits for subcategories of the electroplating industry, including common metals, anodizing, coatings, chemical etching and milling, electroless plating, and printed circuit board manufacturing. The limits for precious metals plating are similar to those shown in Table 26.11. However, precious metals plating also includes limits for silver dischargers that discharge more than 38,000 l/day of treated wastewater. These silver limits are a one-day maximum of 1.0 to 2.0 mg/l and a four-day consecutive average of 0.7 mg/l. Wastewater treatment sludges from electroplating operations are regulated as hazardous wastes under RCRA in the United States. These sludges are listed hazardous wastes, with the RCRA classification F006. They can contain a variety of metals, including chromium, cadmium, copper, nickel, tin, zinc, and others, as well as metal–cyanide complexes. Most of the sludges are placed in hazardous waste landfills after treatment for stabilization [35,36], which typically involves chemical additions to precipitate and immobilize the metals. Cementitious stabilizing agents are often used. Some electroplating sludges are processed to recover valuable metals. However, the costs of this processing coupled with constraints on recycling in RCRA regulations have tended to favor landfill disposal as the preferred management approach. The electroplating industry has been working with the USEPA for revision of RCRA regulations to create incentives for development and implementation of metal recovery and recycling techniques. A step to promote recovery of metals from electroplating sludges was taken by the USEPA in 2000, when RCRA regulations were revised to

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Soak Clean

Electro Clean

Rinse

Acid

539

Rinse

Ni Plate

Work Flow Rinse

Dry

Hot Rinse

Cr-Plate

Pre-soak

FIGURE 26.8 Typical electroplating line for steel. Courtesy of F. Altmayer, Scientific Control Labs, Inc.

TABLE 26.10 Metals Used in Bath Formulations in Electroplating and Metal Finishing Bath formulations Metals

Cyanide compounds

Cadmium Chromium Copper

X

Brass Bronze Gold Lead and lead/tin Nickel Rhodium Silver Tin

X X X

X

Noncyanide compounds

X X (acid bath) X (phosphate) X (formaldehyde)

X X X X X

Source: Schlesinger, M., Kirk–Othmer Encyclopedia of Chemical Technology, online edition [June 18, 2004], John Wiley & Sons, New York, Copyright 2004. John Wiely and Sons, Inc. Reprinted with permission of John Wiley & Sons, Inc.

extend generator accumulation time limits for F006 electroplating sludge [36]. This ruling allows large quantity generators to accumulate F006 waste on-site for up to 180 days (or 270 days in certain circumstances) without having to obtain a RCRA permit. The intent in this regulatory change was to make it more economically feasible to recover metals from sludge by accumulating larger quantities for shipment and processing.

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TABLE 26.11 Effluent Discharge Limits for Electroplating and Metal Finishing Pollutant

Daily maximum, mg/l

Four day consecutive avg., mg/l

(a) Pretreatment standards for electroplating (40 CFR Part 413) Cadmium 1.2 Lead 0.6 Cyanide (amenable) 5.0 Total toxic organics 4.57 Cadmium 1.2 Chromium 7.0 Copper 4.5 Lead 0.6 Nickel 4.1 Zinc 4.2 Silvera 1.2 Cyanide (total) 1.9 Total toxic organics 2.13

Pollutant

0.7 0.4 2.7 NS 0.7 4.0 2.7 0.4 2.6 2.6 0.7 2.7 NS Max. Monthly Avg, mg/l

Daily maximum, mg/l

(b) Pretreatment standards for metal finishing (40 CFR Part 433) Cadmium 0.69 Chromium 2.77 Copper 3.38 Lead 0.69 Nickel 3.98 Zinc 2.61 Silvera 0.43 Cyanide (total) 1.2 Total toxic organics 1.23

Chromium Cyanide Zinc

0.26 1.71 2.07 0.43 2.38 1.48 0.24 0.65 NS

mg/m2

lbs/million ft2 processed

mg/m2

lbs/million ft2 processed

0.18 0.095 0.49

0.037 0.02 0.10

0.072 0.038 0.20

0.015 0.008 0.041

NS: no limit specified. a Applies only to precious metals plating.

26.2.2.3 Wastewater Management Strategy Most electroplating and metal finishing operations are located within urban settings and wastewaters are generally discharged to a POTW. As such, primary compliance issues are related to the loadings of cyanide (usually specified as total cyanide) and heavy metals in the discharges to the local POTW, and management of the wastewater treatment sludge. Electroplating and metal finishing operations have had to contend with the treatment of cyanide and heavy metal discharges for a long time and a great deal of information on the treatment of these constituents is available [37]. The 1980 review by Zabban and Helwick [38] is still very pertinent and their conclusion that “chlorine and hypochlorite oxidation of cyanide is the most acceptable

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conventional treatment” is still true. Table 26.12 presents a list of the technologies reviewed by Zabban and Helwick along with a list of those reported in the recent literature. Details of the technologies listed in Table 26.12 are discussed in Chapters 20 and 21 Table 26.12 shows conventional technologies for end-of-pipe cyanide treatment. Recent literature reports updated information for some of these technologies. For example, earlier investigations indicated that electrochemical oxidation was effective “only” for high strength cyanide wastewaters (e.g., spent plating baths); investigations that are more recent have concluded that this technology may perform well for the treatment of weaker strength wastewater (e.g., rinse waters). Similarly, improved economics may result in the selection of technologies that offer multiple benefits including cyanide treatment and metals recovery, for example, electrochemical oxidation of cyanide with metals recovery. In addition, it is almost a certainty that as discharge limits become more stringent, end-ofpipe treatment will become more complex and costly. Figure 26.9 presents a generalized end-of-pipe treatment system for the discharge of plating wastewater to a POTW. As shown in the diagram, a two-stage treatment process is typically implemented to provide complete destruction of cyanide to carbon dioxide and nitrogen gas. However, for some POTWs, a single stage oxidation of cyanide to CNO− is acceptable. Cyanate will hydrolyze to CO2 and NH3 , at a rate dependent on pH (see Chapter 5). In response to the movement toward more stringent discharge limits, the long-term economic liability associated with the management of metal-bearing sludges, and the dictates of the Pollution Prevention Act, the emphasis of wastewater management in the electroplating and metal finishing industry has shifted to pollution prevention through recycle and recovery of “waste materials” and process modifications to replace cyanide with nonpolluting chemical reagents. In 1994, the USEPA published a “Guide to Cleaner Technologies and Alternative Metal Finishes” [34]. This guide describes alternative technologies that can be used to reduce wastes and emissions from metal finishing operations. Table 26.13 lists the pollution prevention technologies that were evaluated for electroplating and metal finishing operations. These technologies include both recovery processes as well as processes that use chemical reagents other than cyanide. The examination of these alternative chemical processes are outside the scope of this book. They merit evaluation and consideration, however, when considering the management of cyanide discharges. Many of these alternative processes have been shown to be effective as indicated by the noncyanide plating bath formulations listed in Table 26.10. Commercialization of noncyanide formulations has been limited because of reagent cost and, in some cases, product quality issues, but their development continues. The recovery processes, especially the electrochemical processes, offer opportunities to recover valuable raw materials (i.e., plating metals) while reducing or eliminating the need for end-of-pipe wastewater treatment and hazardous sludge disposal.

26.2.3 BLAST FURNACE REDUCTION OF METAL ORES 26.2.3.1 Process Description Metallic ores (e.g., containing iron and nonferrous metals such as lead and manganese) are thermally reduced to metal in blast furnace processes. For the reduction of iron ore, a mixture of iron ore, coke, and a flux (limestone or dolomite) is fed into the top of the furnace. Heated air and a fuel (gas, oil, or powdered coal) are injected into the bottom of the furnace. The controlled injection of air provides oxygen that partially burns the fuel and coke to produce a high-temperature, reducing environment containing hydrogen and carbon monoxide. Under these conditions, the iron ore (basically iron oxide) is reduced to form liquid metal. This high-temperature, reducing environment also provides the opportunity for the reaction of nitrogen (from the air), carbon monoxide, and hydrocarbons from the injected fuel to form hydrogen cyanide and ammonia. This formation of hydrogen cyanide and other cyanide compounds in the blast furnace is widely recognized. These compounds have long been

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TABLE 26.12 List of Technologies for Treatment of Cyanide in Electroplating and Metal Finishing Wastewaters Zabban and Helwick [38] 1. Alkaline chlorination Effective on WAD cycle Chlorine demand affected by ammonia and organics

2. Ozonation Few ozonation plants Merits additional consideration Studies required to establish control of ozone consumption Ozone demand not affected by ammonia Capital and operating cost greater than that for chlorination 3. Aldehyde and peroxides (e.g., Kastone) Comparable to first stage of chlorination Incomplete oxidation Produces/increases wastewater BOD and TOC

4. Electrochemical oxidation Limited to treatment of high CN concentrations (>100 mg/l) Familiar technology to electroplaters Less expensive than chlorination

5. Precipitation with iron salts May not be a desirable process because of long-term responsibility (RCRA) for disposed materials 6. Adsorption onto activated carbon Is not promising from commercial standpoint

Recent evaluations

High temperature alkaline chlorination treatment of plating sludges (>700 ppm total CN to <500 ppm total CN) [72] (Lancy process)

Cyanide destruction with ozone, Caro’s acid or H2 O2 , high costs [73]

H2 O2 treatment of CR(III) and CN in plating wastes [74] UV–H2 O2 pilot treatment of free and combined cyanides (including Fe and Ni, and organic agents [75] to electroplaters, less expensive than chlorination) H2 O2 with dilute sulfuric acid treatment of nickel cyanide [76]

Electrochemical oxidation of complex cyanide with copper recovery; wastewaters with 250 mg/l CN reduced to 8 mg/l [77] Electrochemical–copper-catalyzed oxidation of dilute (<100 mg/l) cyanide wastewater, achieved less than 1 mg/l CN [78] Treatment of spent bath to remove metal and cyanide [79] Electrochemical destruction of cyanide with zinc recovery [81]

Chemically modified activated carbon improves adsorption capacity for Cu, Cr, Zn, and cyanide from wastewater Cu-catalyzed adsorption and air oxidation of cyanide in GAC adsorption; bench-scale and pilot studies [82] Practiced in gold mining

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TABLE 26.12 Continued Zabban and Helwick [38]

Recent evaluations

7. Ion exchange Concentrating technology Can be used profitably in gold and silver recovery Problematic with other cyanide complexes, questionable practicality 8. Reverse osmosis Concentrating technology Requires effective pretreatment Concentrates need further treatment for disposal 9. Biooxidation No application to plating wastes

Commercial practice in gold mining

10. Acidification and air stripping No applications to plating wastes Venting of stripped CN-rich vapors to the atmosphere prohibited CN recovery may not be economically attractive

Commercial practice in gold mining

NaOH pH ~ 9-10 pH CN rinse Water

NaOCI

Lime

to POTW Clarifier

Two-stage CN destruction H2SO4 Dewater

pH ~ 2 pH Sodium Blsuiflte

Sludge cake to disposal

HexCr Rinse

Hex Cr reduction Other rinse Waters

FIGURE 26.9 Typical electroplating rinse water treatment system.

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TABLE 26.13 Pollution Prevention Technologies Evaluated for Electroplating and Metal Finishing Recovery processes Electrochemical processing Copper recovery with cyanide oxidation Zinc recovery from spent bath with cyanide oxidation Zinc recovery from rinse water, with cyanide oxidation; compared to other treatment methods electrolyte treatment is extremely cost effective Chemical recovery processing Precipitation of silver as AgCl, purification by boiling in 10% HCl solution Silver cyanide precipitation from spent baths Vacuum evaporation Recovery of dilute CN-containing rinse waters

CN processing alternatives

Alkaline cyanide-free galvanizing

Electrodeposition of copper from sulphamate and acid sulfate electrolyes Gold plating from mercapto complexes

Phosphate alkaline copper electroplating Perchlorate zinc electroplating Alkaline Cu-plating process Alkaline Zn and Zn–K combinations Fluoborate and perchlorate based Cd plating Simulated brass (acid–chloride based) process

Source: Data from Slawski, K., Wedzicha, L., and Mromlinska, Z., Rudy I Metale Niezelazne, 37, 248, 1992.

detected in furnace off-gas and residue. In 1837, potassium cyanide deposits were discovered around the injection orifices of furnaces. These deposits contained as much as 43% potassium cyanide, and at that time they were used as a commercial source of cyanide [39]. Blast furnaces are operated on a routine schedule of batch “charges” consisting of a predetermined sequence of ore, coke, and limestone. Molten iron is intermittently removed from the bottom of the furnace as part of the process known as “tapping.” The tapping frequency can vary from 4 to 10 times per 24-h day. Figure 26.10 presents a typical blast furnace material balance for the production of one short-ton of iron [40]. This figure shows that 63,500 scf of off-gas is produced for every ton of iron produced. Blast furnace off-gas, rich in carbon monoxide and hydrogen and with trace amounts of ammonia and cyanide, is used as fuel within the steel mill. Prior to this use, the gas, which contains large quantities of entrained dust and particulate matter, must be cooled and cleaned. The cooling and cleaning of this gas is accomplished by applying the following processing equipment: (a) dry dust catcher, (b) wet scrubber, (c) direct contact water cooling, and (d) electrostatic precipitation. The dry dust catcher removes about 60% of the entrained particulate matter and the latter three operations essentially remove the remaining 40%. These latter stages use water in varying amounts and it is this water that cools the gas and becomes the blast furnace gas scrubber effluent. In the past, blast furnace gas scrubber water was used on a once-through basis, but recirculation systems have replaced the once-through systems. Figure 26.11 presents a schematic flow diagram of a typical recirculation blast furnace gas scrubber system with material balance for a 3300 ton/day blast furnace operation. The figure shows that 2077 gallons of water/ton of iron produced is used for cooling and cleaning the off-gas, and 125 gallons of wastewater/ton of iron is produced. Table 26.14 shows the chemical constituents, for example, ammonia, cyanides, phenol, and suspended solids, in blast furnace gas scrubber effluent [41] from data provided by 19 different operations. These data show variations inherent in processing materials of variable geologic composition and variations in furnace operations.

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INPUTS IRON BEARING BURDEN

Iron Ore Fluxed Sinter Scrap

FLUX

Limestone Gravel

FUEL

Coke

OUTPUTS 615 lb 2452 lb 197 lb

Top Gas 63,500 scf Moisture 3,300 scf Dust & Sludge

4921 lb 157 lb 84 lb

15 lb 15 lb 1028 lb

Slag

BLAST

Air Moisture

FUEL

Natural Gas

44,280 scf 670 scf

3277 lb 32 lb

962 scf

41 lb

498 lb

Hot Metal 2000 lb 4.1%C, 0.90% Si, 0.35% Mn, 0.026% S, 0.296% P Runner Scrap

12 lb

FIGURE 26.10 Typical blast furnace material balance for iron making. All quantities in amount per ton of hot metal. (Source: USS, The Making, Shaping, and Treating of Steel, 9th ed, McGannon, H.E., Ed., US Steel Corp., Pittsburgh, PA, 1971. With permission.)

26.2.3.2 Regulatory Discharge Requirements Discharge of blast furnace scrubber water in the United States is subject to categorical industrial effluent discharge limits, which are based on the BAT technologies (see Chapter 18). Categorical industrial effluent discharge limits for iron and steel blast furnace operations are presented in Table 26.15; included in this table is an effluent limit for cyanide. Although the cyanide species is not specified, regulators tend to assume “total cyanide”. In some instances, when the water quality of the receiving stream exceeds state or federal water quality criteria for specific pollutants, the discharger of a blast furnace scrubber water, or any other point source discharger, may be required to comply with discharge limits that are more stringent than the categorical industrial effluent standards. These discharge limits would be determined based on a site-specific assessment of the affected receiving water body [42]. 26.2.3.3 Wastewater Management Strategy Effective management of blast furnace scrubber water effluent consists of several components: (1) limiting the transfer of constituents of concern from the blast furnace off-gas to the gas scrubber water, (2) separation of the suspended solids from the scrubber water, (3) recycle of the treated scrubber water to the blast furnace gas scrubber, and (4) end-of-pipe treatment system for the blowdown from the scrubber water recycle system. 26.2.3.3.1 Limiting constituent transfer to scrubber water The regulated constituents in blast furnace scrubber water typically are ammonia, cyanide, phenol, and suspended solids. The source of the phenol is generally believed to be in the coke feedstock to the

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Gas In

Gas Out Gas = 14,700 lb/min Dust = 77.0 lb/min Water = 44 gpm Temp = 400° F

Gas = 14,700 lb/min Dust = 7.0 lb/min Water = 132 gpm Temp = 110° F

1490 gpm

Venturi Scrubber

Water in = 4,760 gpm in NH3 = 125 mg/L CNT = 2.5 mg/L OOH = 1.0 mg/L S.S = 30-50 mg/L pH = 7.0-7.5 Temp = 100° F

Direct-Contact Cooler 207° F

110° F 1,402 gpm 174° F

110° F

3,270 gpm Water in=4,672 gpm NH3 =135 mg/L CNT =2.7 mg/L OOH =1.1 mg/L S.S =1000-3000 mg/L pH =7.0–7.5 Temp =130° F

Controller pH

Polyelectrolyte

4,787 gpm

Betz 3300 40lb/Day

H2SO4 100 Gal 70%/Day

138 gpm

115 gpm

Anti Scale 3 Gal/Day

Clarifier 110⬘ O Res. Time 4.5 Hrs

29,700 Gal Hot Well 33' x 12' x 10'

Make Up H2O

114° F

100° F

Cooling Tower 1868 ft 2

397 gpm Sludge to Sinter Plant Vacuum Filter 930 ft2

172 lb Wet/Min 50% Moisture Fe ~ 36% CaO ~ 6.0 MgO ~ 2.5%

Equivalent to 125 Gal/Ton

Dry Basis

NH3 < 1.0 mg/L CNT < 0.1 mg/L OOH < 0.01 mg/L S.S < 10 mg/L pH ~ 7.0 - 8.0

Blowdown* Flow = 286 gpm NH3 = 125 mg/L CNT = 2.5 mg/L OOH = 1.0 mg/L S.S = 30-50 mg/L pH = 7.0-7.5 Temp = 100° F

FIGURE 26.11 Schematic flow diagram of process profile for model 3300 ton/day blast furnace gas scrubber and cooling waste water treatment facility. (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute report to American Iron and Steel Institute, 1976. With permission.)

blast furnace, especially when this coke has been quenched with water containing coke plant flushing liquor. This practice is called “dirty quenching” because of the contaminants that are present in the flushing liquor, one of which is phenol. In U.S. plants, this dirty quenching is no longer practiced or allowed.

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TABLE 26.14 Composition of Clarified Blast Furnace Gas Scrubber and Direct Cooling Water Cyanides Ammonia

Total

Free

Phenol mg/l

Systema

pH

mg/l

6–7

3–12 18–28

0.01–0.2 0.1–0.8

0.1–0.3 0.1–0.8

2 3

OT RS RS RS

143±35

10.7

1.2±1.5

4

OT

5 6 7

OT OT RS

9

OT

10 11 12 13A 13B 14

OT RS OT OT OT RS

3.18–8.91 5.73 0.80 2.58 [465] 135 0.025–0.15 0.115 0.3–4.8 0.06–8.3

0.02–0.04 0.03 0.32 0.02 [4.5] 2.5 0.14–0.32 0.22 0.05 0.01–0.07 0.7–1.2

15 Interlake (5) YS&T (6) 16A 16B 17

RS RS RS OT OT RS

Company 1

7.5 7.3–8.0 7.5 7.1–7.8 7.5

14.4 6.46 [734] 150 1.7–4.1

mg/l

8.5 [9.2]b 8.5 8.0–9.6 8.9 6.5–8.5 7.4–8.6 7.0

0.03–0.15 8.6–43.7 2–15

6.8–7.9 7.5 8.4 8.5 7.8–8.6

40–150 75 146 9.1 186

0.002–5.0 0.26 6.02 0.30 0.73

81

3.25

[18] 3.98

4–20

5.2

1.24 0.54 14.6

0.02

Susp. solids mg/l

Clarif. Overflow gpm/ftb

Floc. agent

65 82 49–77 69±40

0.56

Yes

0.55 0.61

No Yes

34–125 52 40 20 [597] 170 22–74 48 200 16–394 55 27.0 41.5 20–100 52 44 80 90 32±24 64±64 54

1.99

No

0.80 0.32 1.83

No Yes Yes

0.80

Yes

0.44 0.31 0.33 0.825 0.465 1.46

No No No Yes No Yes

0.51 0.79 0.76 0.33 0.19 0.71

No No No

Yes

a OT = once through, RS = recycle system. b [] indicate max. observed values.

Source: Data from Wong-Chong, G.M. and Caruso, S.C., An evaluation of the treatment and control technology recommended for blast furnace (iron) wastewater, Carnegie Mellon Research Institute report to American Institute of Iron and Steel, Pittsburgh, PA, 1976.

The ammonia and cyanide produced in the blast furnace end up in the blast furnace off-gas and are subsequently transferred to the gas scrubber water. The following discussion on the transfer of these constituents, and phenol, was taken from a 1976 Carnegie Mellon Research Institute report to the American Iron and Steel Institute [41]. During the scrubbing of the blast furnace gas, the mass transfer of ammonia, hydrogen cyanide, phenol, and carbon dioxide can be described with a Fickian mass transfer expression: dN/dt = −DA[dC/dx]

(26.3)

where N is the moles of compounds transferred to liquid phase mol/h, D the diffusion (mass transfer) coefficient, ft2 /h, A the interfacial area, ft2 , C the concentration of compound A, mol/ft3 , x the distance measured in direction of diffusive transport, ft, and t the time, h.

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TABLE 26.15 Effluent Discharge Limits for Blast Furnace Wastewater (40CFR Parts 420.33, 420.34 and 420.35) BAT, PSES, and NSPS discharge limits, kg/kg product Parameter Ammonia–N CN (unspecified) Phenol Lead Zinc TRCb TSSa O&Ga

1-day maximum

30-day average

0.00876 0.00175 (0.000584)a 0.0000584 0.000263 0.000394 0.000146 0.0117 0.0029

0.00292 0.000876 (0.000292)a 0.0000292 0.0000876 0.000131 NSc 0.00438 NSc

a Specified for NSPS in addition to other parameter limits. b Specified for BAT and NSPS. c NS = no limit specified.

The concentrations in the liquid and gas phase are influenced by temperature and pH; mass transfer profiles are shown pictorially in Figure 26.12. For a given set of operating conditions, the mass transfer will eventually reach a steady state condition. At this time, the bulk liquid concentration will not change with time. However, changes in pH and other scrubber water quality parameters can shift the equilibrium such that changes in the bulk liquid concentration will occur. The directions of this change in composition for each component can be predicted from an examination of the following aqueous reactions: Ammonia :

− NH3 + H2 O
NH+ 4 + OH

Carbon dioxide : Hydrogen cyanide : Phenol :

2CO2 + 2H2 O
H2 CO3 + 2H+ + CO− 3 −

HCN
CN + H

+

C6 H5 OH
C6 H5 O− + H+

(26.4) (26.5) (26.6) (26.7)

Under acid conditions (i.e., elevated concentrations of H+ and low concentrations of OH− ) reactions (26.5) to (26.7) will shift to the left and reaction (26.4) to the right, increasing the concentrations of HCN, phenol, carbon dioxide, and the ammonium ion in solution. Because phenol, carbon dioxide, and hydrogen cyanide are volatile species, their potential to return to the gas phase is much greater than the dissociated, NH+ 4 ion. This results in reduced removal of CO2 , HCN, and phenol from the off-gas. However, at the same time, the removal of ammonia, NH3 , will be high, as it will tend to dissolve to form NH+ 4 . In contrast, under alkaline conditions, transfer from off-gas to scrubber water will be low for ammonia but high for CO2 , HCN, and phenol. Monitoring of ammonia, cyanide, and phenol concentrations, and pH in the blast furnace scrubber water at an operating facility (Company 1) provided the data shown in Figure 26.13. Both the phenol and cyanide concentrations increased significantly with increase in pH. While these data were from short-term observations, long-term operation data from another facility (Company 3) showed

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Gas bulk

Gas film

Liquid film

CO2 NH3 in liquid phase HCN øOH

Liquid bulk

Pi

P Pi

Ci C

Concentration of

Partial pressure of

CO2 NH3 in gas phase HCN øOH

Interface

C

Direction of mass transfer

FIGURE 26.12

Illustration of mass transfer of HCN, CO2 , NH3 , and phenol from the gas to the liquid phase.

11.7 CN

Phenol and CN concentration, mg/L

2 CN

Phenol 1

0 4

5

6

7

8

pH

FIGURE 26.13 Relationship of cyanide and phenol concentrations to pH in blast furnace scrubber water. (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute report to American Iron and Steel Institute, 1976. With permission.)

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Cyanide concentration, mg/L

100

Ln CN = 1.53 Ln ALK–8.1 CN = ALK1.53e–8.1 10.0

1.0

0.1 10

100

1,000

10,000

Alkalinity, mg/L

FIGURE 26.14 Effect of alkalinity on cyanide transfer in blast furnace scrubber water. (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute report to American Iron and Steel Institute, 1976. With permission.)

a similar relationship between pH and CN concentrations in the scrubber water, and also showed the influence of alkalinity on the transfer of HCN from the gas to the aqueous phase. Figure 26.14 shows the relationship of scrubber water cyanide concentration and alkalinity at a pH of 8.3, which can be described mathematically as: CNT = ALK1.53 e−8.1

(26.8)

where CNT is the total cyanide concentration in scrubber water, mg/l as CN, and ALK the alkalinity in the scrubber water, mg/l as CaCO3 . Similar relationships were developed for pH levels 7.8, 7.9, and 8.0. Since pH and alkalinity are related, the data in Figure 26.14 were further examined by plotting the cyanide concentrations in the scrubber water against the product of pH and alkalinity (Figure 26.15). Similar data were also examined to determine the effects of pH on ammonia concentrations in the scrubber water. Figure 26.16 presents these data, which indicate, as expected, that ammonia concentrations in scrubber water decrease as pH is increased. This is opposite of the behavior of cyanide and phenol. From the above analysis, it is apparent that the transfer of the constituents of concern — ammonia, cyanide, and phenol — can be controlled by controlling the pH or alkalinity of the blast furnace gas scrubber water. However, pH has different effects on the transfer of cyanide and phenol, and ammonia: low pH conditions favor the transfer of ammonia to the aqueous phase while high pH conditions favor the transfer of cyanide and phenol. An optimum pH condition can be determined from an assessment of the costs associated with the treatment of these constituents in the

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Average CN concentration, mg/L

100

10.0

1.0

0.1 100

1,000

10,000

100,000

Product of pH and alkalinity: (pH) (ALK)

FIGURE 26.15 Relationship between pH, alkalinity, and cyanide concentrations in blast furnace scrubber water (Data from Company 3). (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute report to American Iron and Steel Institute, 1976. With permission.)

blowdown from the blast furnace gas scrubber system and the cyanide generating characteristics of the furnace. 26.2.3.3.2 Blowdown treatment Table 26.14 presents data on blast furnace gas scrubber water blowdown from several plants. Overall, these data show significant variations in the concentrations of ammonia (0.03 to 743 mg/l), total cyanide (0.002 to 405 mg/l), phenol (0.01 to 14.6 mg/l), pH (6.0 to 9.6), and suspended solids (16 to 597 mg/l) as well as pH (6.0 to 9.6). These variations are due to the specific operating characteristics of each facility and the inherent compositional variations in the iron ore processed. The operating characteristics include the quality of the scrubber water (especially pH and alkalinity), the number of cycles of recirculation, the overflow rate and condition of the scrubber water clarifier, and the use of flocculating agents. Figure 26.3 presents a general listing of available technologies for the treatment of ammonia, cyanide, phenol, and suspended solids; phenol is represented by the line(s) for “organics.” These technologies can be applied to the treatment of blast furnace scrubber water blowdown. For cyanide, the treatment selection will depend on the cyanide species that are present: free or weak acid dissociable vs. strong metal–cyanide complexes. The data in Table 26.14 suggest that the dominant cyanide species are the metal–cyanide complexes, that is, 18 mg/l free vs. 465 mg/l total cyanide, and 4 mg/l free cyanide vs. 135 mg/l total cyanide. However, in one case the free cyanide concentration was comparable to the total cyanide concentration, that is, 5.2 vs. 6.0 mg/l. If the metal–cyanide complexes are dominant, it is likely that they will consist mostly of the iron–cyanide complex because

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200

180

Ammonia Concentration, mg/L

160 140

120 100

80

60

40

20 5.5

6.0

6.5

7.0

7.5

8.0

8.5

9.0

pH

FIGURE 26.16 Effect of pH on the ammonia concentration in blast furnace scrubber water (Data from Company 3). (Source: Wong-Chong, G.M. and Caruso, S.C., Carnegie Mellon Research Institute report to American Iron and Steel Institute, 1976. With permission.)

of the large quantity of iron that is present in the scrubber water and the high affinity of cyanide for iron (see Chapter 5). This is indicated by the iron content of the scrubber water clarifier sludge cake, about 36% iron on a dry weight basis (see Figure 26.11). For blast furnace scrubber water blowdown containing ammonia, iron–cyanide complexes (predominant cyanide species), WAD cyanide, and phenol, two treatment trains have been shown to provide effective treatment: • Treatment Train 1 (Figure 26.17): Fixed-film biological treatment of ammonia, WAD cyanide, and phenol preceded by chemical precipitation of iron–cyanide complexes. Constituent concentrations in the effluent from this process of <3 mg/l ammonia, <1 mg/l total cyanide, and <0.1 phenol can be expected. • Treatment Train 2 (Figure 26.18): Physical–chemical treatment of all parameters of concern can be achieved using precipitation followed by chlorination. The precipitation removes the iron–cyanide complexes and WAD cyanide; the chlorination removes ammonia, WAD cyanide, and phenol. Effluent quality can be similar or better than that achieved in Treatment Train 1, depending upon the chemical dosages that are used. Generally, where applicable, biological treatment tends to be more economical than physical– chemical treatment options. Furthermore, the reduced chemical usage in biological treatment options tends to result in a better quality effluent, that is, lower total dissolved solids, which can result in improved toxicity characteristics based on whole effluent toxicity testing.

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15:1 Fe\CN

Sludge disposal

Sludge dewatering

Cyanide reaction Tank

6.0 - 6.5 Ferrous sulfate

Lamella clarifier

Air

Fixed film biological tower

pH

Alkali (NaOH)

NH3 = < 3 mg/L CNT = < 1 mg/L Phenol < 0.1 mg/L Nitrate - High* TSS, 20 mg/L

Discharge

Nitrate - N = (Input - Effluent) Ammonia-N

FIGURE 26.17 Treatment train no.1: chemical/biological treatment of blast furnace gas scrubber blowdown.

Ammonia Cyanide Phenol Susp. Solids

Scrubber water blowdown

Acid (H2SO4)

pH

Management of Cyanide in Industrial Process Wastewaters 553

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Sludge

Cyanide reaction tank

pH

15:1 Fe\CN Ferrous sulfate

Lamella clarifier

Chlorine or NaOCl

Breakpoint chlorination

FIGURE 26.18 Treatment train no.2: chemical treatment of blast furnace gas scrubber blowdown.

Susp. solids

Phenol

Cyanide

Ammonia

Scrubber water blowdown

Acid (HSO)

6.0 - 6.5

ORP

Bisulfide

Dechlorination

Discharge NH3 = < 3 mg/L CNT = < 1 mg/L Phenol < 0.1 mg/L TSS, 20 mg/L

ORP

554 Cyanide in Water and Soil

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26.2.4 COAL COKING 26.2.4.1 Process Description Coke is produced by the destructive distillation of coal in the absence of air. This process operates at temperatures as high as 1100◦ C (2000◦ F) and drives off volatile components including tar, organic compounds, moisture, and water of combustion from the coal [39]. Tar, water and organic vapors, and other gases are carried from the ovens by means of reduced pressure and condensed in the gas mains and primary cooler. Continuous decanters separate the tar from the condensed aqueous materials, which are called flushing liquor or waste ammonia liquor (WAL). In the coking process, cyanogenic compounds are produced by the reaction between organic hydrocarbons (e.g., methane) and nitrogen under reducing conditions and high temperature (>1000◦ C). Nitrogen is available from air trapped in the initial coal charge and nitrogenous compounds contained in the coal. About 1.5 to 2.0% of the coal nitrogen is converted to cyanogenic compounds [39]. These reactions would be similar to those in the experimental reactions of Stanley Miller discussed in Chapter 1. Detailed discussions of the theories regarding cyanogenic compound formation and reactions that may produce them are presented by Robine and Lenglen [39]. Table 26.16 presents a qualitative composition of coke oven gas (COG). Trace constituents include hydrogen cyanide, methylcyanide, and thiocyanate. On cooling of the COG, a portion of these constituents partition to the condensate; this condensate is called Waste Ammonia Liquor (WAL). Figure 26.19 presents a schematic flow diagram of a semidirect (gas cooling by both direct WAL injection and indirect cooling) by-product coke plant. As shown, a major part of the plant is dedicated to COG cleaning. In many modern coke plants, the COG leaving the gas scrubber is further processed for sulfur removal. Figure 26.19 also shows the four primary wastewater streams that are produced in a by-product coke plant: (a) WAL; (b) final cooler blowdown; (c) benzol plant blowdown; and (d) ammonium sulfate crystallizer blowdown. Other wastewaters include blowdown from the tar still, and drip leg condensate from the gas mains and COG desulfurization. The two streams from the tar still and drip legs are relatively small, intermittent, and have composition very similar to WAL. Table 26.17 presents a detailed composition of WAL. These data show cyanide, as HCN, in the 50 to 62 mg/l range and ferrocyanide in the 14 to 39 mg/l range [41,43]. Table 26.18 presents summary characteristics of WAL from 11 coke plants. These data along with those in Table 26.16 show considerable variations in both chemical composition and quantity of WAL produced. This variability is understandable given the variability in coal composition from one mine to another and water content of the coal when charged to the ovens; stock piled coal could be very wet from rainfall or snow. Table 26.19 shows ammonia, cyanide, and phenol content of the major wastewater streams in a coke plant. As indicated there, final cooler blowdown and desulfurizer wastewaters can contain high concentrations of cyanide. Table 26.19 also shows nominal flow rates for each wastewater stream. 26.2.4.2 Regulatory Discharge Requirements Coke plants in the United States must comply with BAT for discharges to receiving streams. These discharge requirements are listed in Table 26.20. The limits given are proposed limits in the 2002 Final Rule (40 CFR Part 420.13) and are based on an average wastewater flow of 163 gallons per ton (gpt) of coke produced as shown in Table 26.21. They include a daily average limit of 3.06 mg/l and a daily maximum of 4.37 mg/l for total cyanide. Coke plant discharges in China and Brazil are being required to meet a more stringent 1.0 mg/l total cyanide. The average 163 gpt flow rate comprises contributions from the various operations within a coke plant and includes a 50 gpt service water addition allocation to assist in the control of the operation of the biological treatment process. It must also be recognized that the discharge of treated effluents to U.S. receiving waters can be dictated by water quality criteria based on designated water use classifications. More stringent discharge quality requirements can result from the use of these water quality criteria.

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TABLE 26.16 Qualitative Composition of Coke Oven Gas Organic compounds Acetylene Ethylene Propylene Butylene Allylene Crotonylene Terrene Benzol Thiophene Styrolene Naphthalene Methylnaphthalene Acetylnaphthalene Fluorene Fluoranthane Methane Propane Butane Inorganic compounds Hydrogen Carbon monoxide Carbonic acid Ammonia Cyanogen Hydrogen thiocyanate Methylcyanide Hydrogen sulfide Carbon disulfide

C2 H2 C2 H4 C3 H6 C4 H8 C3 H4 C4 H6 C5 H8 C8 H6 C4 H4 S C8 H8 C10 H8 C11 H10 C12 H10 C16 H10 C18 H10 CH4 C3 H8 C4 H10 H2 CO H2 CO3 NH3 C2 N2 HCNS C2 H3 N H2 S CS2

Source: Data from Robine, R. and Lenglen, M., The Cyanide Industry, John Wiley & Sons, New York, 1906.

26.2.4.3 Wastewater Management Strategy In the early days, ca. 1900, it was profitable to recover cyanide from coking operations. Robine and Lenglen [39] identified nine processes for the recovery from COG and 16 processes for recovery from WAL. Up to the mid-1960s, economic conditions existed that favored recovery of by-products such as ammonia (as ammonium sulfate) and phenolic coal tar acids. Today these economic conditions no longer exist and recovery is not economical. However, in many plants, ammonia is still recovered as ammonium sulfate or anhydrous ammonia at a significant cost burden. Most coke plants in the United States are located where water is abundant. Therefore, there is little incentive for water recirculation and recovery, and “dirty water” (WAL) coke quenching is now prohibited. Treated wastewater is also considered “dirty water.” The overall quality of coke plant wastewaters, even after treatment, is poor because of high total dissolved solids, which can be greater than 15,000 mg/l. Therefore, economic recovery and reuse of this water is, at best, questionable. For these reasons, coke plant wastewater management essentially is an end-of-pipe activity focused on discharge to local receiving waters. The authors are aware of only one coke plant in the world where water supply is an issue: Tai Yuan Iron and Steel Company, Tai Yuan, China. Implementation

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Gas

a

Waste ammonia liquor storage

Waste ammonia liquor

Decanter

Flushing liquor

To river

Tar

Tar separator

Primary cooler

Tar to storage

Gas

From river

Gas

Tar

Exhauster Tar

(NH 4) 2SO 4

Ammonia absorber

Free leg

Ammonia still

Lime

NH3

Steam

Fixed leg

Crystallizer effluent d Ammonia

Gas

Lime reservoir

Tar extractor

Gas

Still waste

H2SO4

Cooling water

Cooling tower

Air

b

Gas scrubber

Naphthalene slurry

Naphthalene skimmer basin

Final cooling system blowdown

Water

Final cooler

Gas

To gas storage

c

Benzol plant blowdown

Benzol Plant

Wash oil

By-Products

Wash oil from benzol plant

FIGURE 26.19 Flow diagram of a semidirect by-product coke plant. (Source: USS, The Making, Shaping, and Treating of Steel, 9th ed, McGannon, H.E. Ed., US Steel Corp, Pittsburgh, PA, 1971. With permission.)

Coke

Ovens

Coal

Flushing liquid

Cooling water

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TABLE 26.17 Characteristics of Weak Ammonia Liquor in Two Coke Production Plants Plant Parameter

A (g/l)

B (g/l)

Total ammonia Free ammonia Fixed ammonia Carbon dioxide as CO2 Hydrogen sulfide as H2 S Thiosulfate as H2 S2 O5 Cyanide as HCN Thiocyanate as HCNS Ferrocyanide as (NH4 )2 Fe(CN)6 Total sulfur Phenols as C6 H5 OH

7.60 4.20 3.40 2.35 0.86 0.022 0.062 0.36 0.014 1.014 0.66

6.20 4.76 1.44 3.94 0.34 0.51 0.05 0.42 0.039 0.57 3.07

Source: Data from Wilson, P.J. and Wells, J.H., Coal, Coke and Coal Chemicals, McGraw-Hill, New York, 1950.

TABLE 26.18 Summary Composition of Weak Ammonia Liquor from 11 Coke Production Plants Waste generated total WAL gal/ton coke Company 1 2 3 4 5 6a 7 8 9 10 11

Maximum

Average

pH

Total ammonia mg/l

66.0 89.5 72.6 465.5 83.0 82.3 150.0 60.1 100.0 NAb

42.0 46.0 38.4 39.0 40.0 36.0 47.0 30.0 33.0 NA NA

8.7 9.1 NA NA 8.5 6.5–8.5 7.5–8.5 5.5 NA NA 8.8–9.1

5500 2800 9841 6500 5000 1500 3900 2500 NA 3010 1713–3417

Total cyanides mg/l

Phenol mg/l

100 140 10 65 50 15(1400)a 10–100 4 NA NA 10–200

3000 400 1753 1690 2500 550 200–300 200 350 770 660–840

a The value in parenthesis is thiocyanate. b NA: no data available.

Source: Data from Wong-Chong, G.M. and Caruso, S.C., An evaluation of the treatment and control technology recommended for coke plant wastewater, Carnegie Mellon Research Institute report to American Institute of Iron and Steel, Pittsburgh, PA, 1976.

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TABLE 26.19 Composition of Major Wastewater Streams in Coke Plant Operations

Source Tar still WAL Final cooler blowdown NH3 crystallizer Benzol plant Desulfizer

Composition (mg/l)

Flow contribution gal/ton

Ammonia

Cyanide

Phenol

2 35 10 35 23 14

5000 6000 200 11 5 0

50 60 500 15 5 1130

4950 1455 240 <10 10 0

Source: Data from Wong-Chong, G.M. and Caruso, S.C., Advanced biological oxidation of coke plant wastewaters for the removal of nitrogen compounds, Carnegie Mellon Research Institute report to the American Iron and Steel Institute, Pittsburgh, PA, 1977.

TABLE 26.20 Coke Plant Effluent Discharge Limitation Guidelines, 40CFR Part 420.13 Proposed Revision Discharge load

Regulated parameter Ammonia–N Benzo(a)pyrene Cyanide Naphthalene Phenols (4AAP)

Concentrations

Max. daily (lbs/thousand lb of product)

Max. monthly avg. (lbs/thousand lb of product)

Base flow (gpt)

Effluent monthly avg. (mg/l)

Max. daily mg/l

0.00293 0.000011 0.000297 0.0000111 0.0000381

0.00202 0.00000612 0.00208 0.00000616 0.0000238

163 163 163 163 163

2.97 0.009 3.06 0.009 0.035

4.31 0.016 4.37 0.016 0.056

of major water recovery and recycle is planned for this integrated iron and steel manufacturing complex, including recovery of water from the by-product coking operation. Figure 26.20 presents a schematic flow diagram of a typical U.S. coke plant wastewater treatment system that complies with BAT effluent limits. This process essentially consists of the following process stages: • Storage and equalization of the wastewater with very high concentrations of ammonia, cyanide, and sulfide (i.e., WAL, drip leg condensate, tar still blowdown, and final cooler blowdown). • Free/fixed leg steam stripping to remove ammonia, free cyanide, and sulfide. • Equalization and storage of steam stripped wastewaters and other less contaminated wastewaters. • Cooling of wastewater mixture charged to biological treatment to maintain 30◦ C maximum temperature in the aeration basins. • Activated sludge biological treatment with nutrient addition (phosphoric acid) and pH control (6.8 to 7.5) followed by clarification. • Dewatering of excess sludge.

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TABLE 26.21 Average Wastewater Production in Coke Plants, 40CFR Part 420.13 Proposed Revision Process wastewater flow component Base flows applicable to all plants Waste ammonia liquor Crude light oil recovery Final gas cooler condensate Coke oven gas condensate Barometric condenser blowdown Steam/caustic for ammonia still Miscellaneous NESHAPs controls Base flow Control water — biotreatment Base flow with control water Optimal flows up to maximum amounts shown Wet coke oven gas desulfurization Indirect ammonia recovery Unregulated WAPC flows Coke plant groundwater remediation Process area storm water

Waste production rate (gal/ton) 1982 regulation

2002 final rule

Iron & Steel 32 25 10 Not considered 3 13 20 Not considered 103 50 153

Merchant 36 28 12 Not considered 5 15 25 Not considered 120 50 170

All coke plants 32 25 10 3 3 10 20 10 113 50 163

25 60 Not considered Not considered Not considered

25 60 Not considered Not considered Not considered

15 NA Design basis Design basis Design basis

NA = no data available.

As indicated by this sequence, pretreatment of coke plant wastewaters, especially for removal of ammonia and cyanide, is usually implemented prior to biological treatment [44–47]. Ammonia stripping typically reduces ammonia concentrations to a range of 20–100 mg/l as N, and total cyanide concentrations to 2–15 mg/l as CN. Coke plant wastewaters typically have high levels of thiocyanate (SCN− ), with 200–600 mg/l present in the effluent from the ammonia stripping treatment [43–45,47,48]. Figure 26.20 shows that wastewaters from the various sources are combined and directed to a storage/equalization tank. The blended wastewater from storage/equalization tank is then heated and passed through two ammonia distillation units in which ammonia is stripped from the aqueous phase and into the gas phase. The first distillation unit is intended for removal of free ammonia, that is, that fraction of the total ammonia that can be stripped from solution without adding caustic as well as the acid gases hydrogen sulfide (H2 S) and hydrogen cyanide (HCN). Cyanide present in the form of cyanide anion, metal–cyanide complexes, or thiocyanate is not removed in this process. It has been observed that part of the free cyanide present in the influent to equalization/storage can be converted to thiocyanate during storage and distillation, depending on the availability of polysulfide and other sulfur compounds in the water [44,49]. A second distillation unit is employed for removal of fixed ammonia. Caustic is added to raise the pH of the wastewater prior to the second distillation step. Wastewater exiting the ammonia stripping process is then mixed with some smaller volume process waters in an equalization tank, and directed to a closed-circuit cooler. Chemical additions (phosphate, lime) are made to the flow of ammonia-stripped wastewater for nutrient addition and pH adjustment prior to biological treatment. The wastewater flow is then directed to activated-sludge basins for removal of phenolic compounds, cyanide compounds, and the remaining ammonia. Dilution water, consisting of supply water or heated water from process cooling, is sometimes added to activated sludge unit influent for dilution of the chemical concentrations and hence toxicity in the influent [45].

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Scrubbers on pushing emission control

To tar decanters

Tar

Steam

NaOH

Equalization tank

Blowdown to quench station

Equalization tank

Cooler

Air

Single stage aeration basin

Alkalinity/pH control

Nutrient addition

Return sludge

Sludge cake to coke ovens

Clarifier

Treated effluent

FIGURE 26.20 Schematic flow diagram of BAT treatment for by-product coke plant wastewater. (Adapted from: Baker, J.E. and Thompson, R.J., EPA-R2-73-167, USEPA, 1973.)

Ammonia sulfate crystallizer blowdown

Misc. process

Benzol plant blowdown

Tar still blowdown

Waste ammonia liquor

Drip leg condensate

Gases to COG main

Free still Fixed still

Final cooler blowdown

Management of Cyanide in Industrial Process Wastewaters 561

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Activated-sludge is employed for oxidation of organic compounds, especially phenol and phenolic compounds, and for oxidation of the ammonia remaining after the ammonia stripping and that liberated from WAD cyanide and thiocyanate degradation. Thus, the bacteria in the reactors include heterotrophic bacteria capable of organic compound oxidation and autotrophic bacteria for ammonia oxidation (referred to as nitrification). Activated sludge biological treatment can reduce WAD cyanide to less than 0.5 mg/l and SCN− to nondetectable or low mg/l levels, though high solids retention times, for example, 40 to 100 days, or high mixed liquor solids concentrations, for example, >5000 mg/l, are needed [44–47,50]. Some iron-complexed cyanide is adsorbed on the mixed liquor biomass. However, at the operating solid retention times (SRT) conditions a major portion of the iron-complexed cyanide will remain in solution and pass through the treatment system, as much as 1 to 2 mg/l depending on the complex cyanide concentration in the raw wastewater. Effluent from the activated sludge units is directed to a clarifier to settle the suspended bacteria, or biosolids. Most of the bacterial biomass is returned to the activated sludge units, though some is dewatered and disposed. The clear water from the top of the clarifiers is discharged. While coke plant biotreatment and other industrial wastewater nitrification systems have been effective in treating waters with relatively high concentrations of ammonia, success depends on an acclimated, healthy nitrifier population, proper use of dilution water, and other factors. It is well known that nitrifying bacteria are slow growing, sensitive species that are susceptible to inhibition from temperature changes and from a variety of toxicants. The potential inhibitory agents for nitrifying bacteria include cyanide and the ammonia itself [51–53; and Chapter 23]. In China and Brazil where stringent cyanide and nitrate discharge requirements are imposed, denitrification and cyanide precipitation treatment are added to the processing train as shown in Figure 26.21. The degree of additional cyanide treatment is controlled by pH and ferrous iron addition (iron to cyanide ratio), as discussed in Chapter 21. At pH of about 6.0–6.5 and iron dosages of about 10:1 iron to cyanide, effluent total cyanide concentrations of about 1–2 mg/l are attainable. It is difficult to achieve less than 1 mg/l total cyanide as the dissolved concentration of cyanide is limited by solubility of the precipitated iron–cyanide solids [54]. Because this precipitation process is a sludge generation process, solids separation can be critical. Entrained suspended solids in the treated effluent could contribute to the overall total cyanide content of the stream and, in some cases, effluent filtration might be required. For the control of effluent nitrate, the Figure 26.21 flow diagram shows multistaged activated sludge biological treatment with downstream denitrification induced by external carbon source (methanol) addition. It is also possible to operate with upfront denitrification followed by downstream nitrification. In this operating mode, the degree of denitrification is controlled by the rate of recirculated mixed liquor from the discharge end of the nitrification stage back to the upfront denitrification stage. In the latter operation, the required external carbon source for denitrification is greatly reduced by using the inherent carbonaceous biological oxygen demand (BOD) of the wastewater. To achieve the desired level of nitrate in the treated effluent, that is, generally about 10 mg/l (i.e., drinking water standard), the nitrifying mixed liquor return to the upfront denitrification stage must be sufficient to dilute the nitrogen concentration in the total incoming raw wastewater to the desired nitrate concentration. For example, assume incoming wastewater nitrogen is 200 mg/l (sum of contributions from ammonia and thiocyanate) and the desired effluent nitrate concentration is 10 mg/l, then the flow of recycled nitrified mixed liquor must be 19 times the flow of the incoming wastewater. This ratio is expressed as: QRN /Q =




 Ni /NO3 − 1

(26.9)

 where Q is the incoming wastewater flow; QRN the return nitrification mixed liquor flow; Ni the total incoming wastewater nitrogen, NH3 + 0.24SCN + 0.54CNF ; NH3 the incoming wastewater

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To coke oven gas line ASW equalization storage tanks (2)

Chemical plant Final cooler Light oil Desulfurizer

Heat exchanger

NH3 still

Biotreatment equalization/ storage tank Service water phosphoric acid acid/caustic antifoam

Flushing liquor

Discharge

Preaeration PAC

To Shanghai Bay Final clarifier

Sludge hopper

Belt filter

Sludge holding tank

Aeration basins

Air

Polymer

Chemical precipitation

Denitrification filter

Chemical Addition FeSO4 H2SO4 NaOH

Clarifier

MeOH

CaCl2 Alum Polymer

FIGURE 26.21 Shanghai, China coke plant wastewater treatment process block diagram. (Adapted from: Wong-Chong, G.M. et al., Proceedings of WEFTEC Latin America, Water Environment Federation, 1998.)

ammonia concentration; SCN the incoming wastewater thiocyanate concentration; CNF the incoming wastewater WAD cyanide concentration; and NO3 the desired effluent nitrate concentration. As illustrated in this example and discussed in Chapter 23, it is necessary to consider the fate of free cyanide and thiocyanate when assessing nitrification of coke plant wastewater, because biological oxidation of cyanide and thiocyanate yields ammonia. In the pH range of 7.0 to 7.5, the reactions are [45,55,56]: + 2− + SCN− + 2O2 + 3H2 O → HCO− 3 + NH4 + SO4 + H

(26.10)

+ CN− + 0.5O2 + 2H2 O → HCO− 3 + NH4

(26.11)

The total ammonia–nitrogen (NH3 –N) input charged to biotreatment can be estimated by [50]: NH3 –N = NH3 + 0.24SCN− + 0.54WAD CN

(26.12)

The biodegradation of thiocyanate is an important factor in the treatment of coke plant wastewaters. Coke plant ammonia stripping unit effluents often contain more nitrogen in the form of thiocyanate than ammonia, for example, in the range of 200 to 600 mg/l thiocyanate [46]. Nitrogen mass balances for activated sludge units at a coke plant [46] revealed that in the absence of substantial nitrification, effluent concentrations of NH3 –N were significantly greater than influent values, primarily from biodegradation of thiocyanate.

26.2.5 PETROLEUM REFINING 26.2.5.1 Refining Process Petroleum refining is the separation and reforming of the hydrocarbon fractions in crude oil to produce specific products such as gasoline, diesel fuel, jet fuel, kerosene, heating oil, lubricating oils,

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tars, asphalt, and coke. The general refining process begins with the removal of water and other gross impurities from the crude oil followed by distillation of the crude oil into various fractions. The heavy, high molecular weight fractions of the crude oil are thermally processed (i.e., thermally cracked) to produce lighter hydrocarbon fractions. Certain intermediate weight hydrocarbon fractions (e.g., naphthas) are also thermally or catalytically cracked to produce the more desirable light hydrocarbon fractions (e.g., gasoline). The high molecular weight residues from the various refining operations are processed to form petroleum coke. Most modern petroleum refineries utilize some form of cracking (i.e., thermal or catalytic) and this process, especially catalytic cracking, is the principal source of cyanide production in a refinery, with some additional contribution from the coking operation [57,58]. Thermal cracking requires temperatures in the range of 455–510◦ C and pressures of 345–2070 kPa to break the heavy petroleum fractions into their lighter components; catalytic cracking requires lower temperatures and pressures [59]. Under these processing conditions, nitrogen- and sulfur-containing materials in the hydrocarbons are broken down to produce ammonia, hydrogen cyanide, and hydrogen sulfide. The cyanide content in the off-gas of the cracking units is typically 5–100 ppm [58]. This twentyfold variation reflects the natural variation in crude oil composition and in the processing conditions of the catalytic reactors. This variability also results in variations in cyanide concentration in the refining wastewater. Figure 26.22 presents a schematic flow diagram of the catalytic cracking process. This flow diagram shows the heavy hydrocarbon and regenerated catalyst entering a reaction chamber, where the high molecular weight hydrocarbons are cracked to produce lower molecular weight hydrocarbons (e.g., gasoline and diesel fuel). The product from this reactor then enters the lower section of a fractionating column where different product fractions of specific molecular weight are recovered. The heaviest hydrocarbon fraction leaves the bottom of the column, and the lightest hydrocarbon fraction, with ammonia, hydrogen cyanide, and hydrogen sulfide, exits the top of the column. The overhead gases are cooled and water is injected into the cooled gas stream. This water removes ammonia, hydrogen cyanide, and hydrogen sulfide from the overhead gas streams, where it is collected in the accumulator or knockout drum. This “sour” water contains 1,000–10,000 mg/l ammonia; 20–200 mg/l total cyanide; 1,000–10,000 mg/l sulfide; and 500–2,000 mg/l phenol. It is sent to a sour water stripper along with other process sour waters [58]. The sour water stripper is a major component of the plant’s wastewater treatment system. 26.2.5.2 Regulatory Discharge Requirements Petroleum refineries in the United States must comply with regulated effluent discharge limits that are defined as part of their NPDES permits. These permits are guided by the USEPA effluent limitation guidelines (ELG) and receiving stream water quality criteria. The ELGs for the petroleum refinery industry (40 CFR Part 419) do not contain limits for cyanide discharges; the parameters of concern are ammonia, BOD, COD, total chromium, hexavalent chromium, oil and grease, phenol, suspended solids, and sulfide. State regulation authorities frequently include cyanide in discharge permits for petroleum refineries, however. For example, because of the growing concern for cyanide discharges into environmentally sensitive receiving waters of the northwest states and coastal waters of California, petroleum refinery dischargers to Puget Sound and to San Francisco Bay have limits for cyanide and have had to conduct various studies of the fate, transport, and effects of cyanide in the receiving waters [60,61] (see Chapter 11). 26.2.5.3 Wastewater Management Strategy Water usage at petroleum refineries is influenced by a number of factors, including ambient temperature effects on cooling tower water losses, and the refining operations employed at a particular facility. Much of the water use is for direct and indirect contact cooling. The water usage at one refinery in Coffeyville Kansas is about 30 gallons of water per barrel of crude processed [62].

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Flue Gas (to final dust collection)

Catalyst regenerator

Cracking reactor

Stripping steam Spent catalyst Regenerated catalysts

Cold water Fractionator

Gas to recovery

Air

Heavy naphtha feed

Sour water Air blower

gasoline light gas oil

Slurry settler

Recycle Heavy gas oil Wash oil Slurry decant oil

FIGURE 26.22 Fluidized-bed catalytic cracking with product separation in a petroleum refinery. (Source: Speight, J., in Kirk–Othmer Encyclopedia of Chemical Technology, John Wiley and Sons, New York, 2000. Copyright © 2004. Reprinted with permission of [Wiley-Liss, Inc. Wiley Publishing, Inc., a subsidiary of] John Wiley & Sons, Inc.)

Figure 26.23 shows a schematic diagram of a petroleum refinery wastewater treatment system. This figure shows that a portion of the treated sour water, following steam stripping, is reused in the crude desalting operation and as wash water. The wash water is then discharged to the plant wastewater treatment system. Other recycle practices for this water include cooling tower make-up and recycle to the catalytic cracker as overhead injection water. Stripped sour water that is not recycled goes directly to the plant biological treatment system prior to discharge. The plant’s biological treatment system is typically an activated sludge system. The purpose of the sour water stripper is to remove the volatile components (e.g., ammonia, cyanide, phenol, and sulfide) from the plant wastewaters. Typical removals that can be achieved are [58]: • 90–98% ammonia • 99% hydrogen sulfide • 20–80% cyanide (depending on species and concentration) In a 1972 study [63], the American Petroleum Institute reported that 37% (industry average) removal of cyanide could be achieved in sour water strippers. Adjustments in pH, stripping rates, and other variations in process parameters had little effect on cyanide removal efficiency. These observations suggest that the residual cyanide in the sour water stripper discharge consists of metal-complexed

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H2O

Vapors to gas plant Cooler Accumulator

Reactor Catalyst recycle

Sour water

To sulfur plant incinerator

Regenerator Sour water stripper

Sour water from other units Fractionating column

Raw oil charge

Sour water stripper bottoms

Fluid catalytic cracker

Crude oil

Desalter

Crude oil To sewer

To sewer

Other plant sources

API separators

Recycle Biological Recycle to to treatment Units for

Recycle to units for washwater

Discharge

FIGURE 26.23 Petroleum refinery wastewater management: treatment and recycle.

cyanide compounds. Iron–cyanide species have been identified [64,65]. More recent investigations suggest that selenocyanate could be a major constituent of these metal–cyanides [66], depending on the geological formation of the crude oil reservoir (e.g., seleniferous marine shale), and that the concentrations of the selenocyanate could be as high as 6 mg/l [67]. When discharge regulations require additional treatment for cyanide removal, a possible treatment option would be copper precipitation of the selenocyanate [64] coupled with iron coprecipitation followed by some form of filtration/adsorption polishing to achieve the desired quality. Achieving treated water quality in the range of 1 µg/l cyanide with precipitation technology is both a technical and an economic challenge (see Chapter 21). A treatment process employing elemental iron for reduction of selenium has been investigated of the bench scale [68]. Biological treatment can also be implemented for selenocyanate [69] and has reportedly been effective in treating selenocyanate in petroleum refinery wastewater due to rapid dissociation of the selenocyanate [70], though no documentation of this was found in the literature.

26.3 SUMMARY AND CONCLUSIONS • Industrial wastewater management, where possible, must consider product recovery, recycle and reuse, waste minimization, and pollution prevention before end-of-pipe treatment. This is an especially important consideration for industries that use cyanide in the production process, such as hydrometallurgical gold mining and electroplating. • The hydrometallurgical gold mining industry has developed and deployed a wide range of technologies to manage its cyanide-containing wastewaters. The scale of cyanide heap

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•

•

•

•

•

567

leaching operations, coupled with variable degrees of commitment to best practices for environmental management in the industry, has resulted in a mixed record of wastewater cyanide control. The industry is working to implement improved recovery, recycle and reuse strategies. The electroplating and metals finishing industry uses cyanide in plating baths for many applications. The industry has an exemplary history of developing recovery, recycle, and reuse as well as alternatives to cyanide. The need for treatment of a wide range of cyanide species in electroplating wastewaters and sludges remains, however. Blast furnace operations in steel making generate cyanide, which is transferred from offgas to gas scrubber water. Effective treatment of off-gas scrubber blowdown usually entails biological treatment for removal of free cyanide and precipitation to remove iron–cyanide complexes. Coke plant wastewater management is essentially an end-of-pipe operation. Treatment usually involves mixing and equalization of four major source flows, followed by steam stripping to remove ammonia and HCN, and then cooling of the water with subsequent aerobic biological treatment to remove remaining ammonia and cyanide. Thiocyanate concentration is high in coke plant wastewaters, and is removed in the biological treatment step. In situations where water recovery is required or effluent limits include dissolved solids, the treatment train would need to include membrane technologies (e.g., ultrafiltration and reverse osmosis or evaporation or distillation) for dissolved solids removal. The petroleum refining industry is increasingly required to control cyanide in wastewater discharges. Conventional chemical oxidation or biological treatment approaches can be used for free and WAD cyanide. Selenocyanate, an unusual cyanide compound, is present in some petroleum refinery wastewaters and treatment could be by copper precipitation. Treatment of industrial wastewaters in the United States must meet federal BAT limits or state-imposed water quality criteria limits based on aquatic toxicity.

REFERENCES 1. Wong-Chong, G.M., Management of industrial wastewaters, in Proceedings WEFTEC Latin America, San Juan, Puerto Rico, 2001. 2. Graves, B.A., Simulated brass process eliminates cyanide, Products Finish., 2, 62, 1997. 3. APHA/AWWA/WEF; Method 4500-CN Cyanide, in Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S., Greenberg, A.E., and Easton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 4. Campbell, J.R., Luthy, R.G., and Dzombak, D.A., Demineralization for reuse of coal conversion condensates, Ind. Eng. Chem. Process Des. Dev., 22, 496, 1983. 5. Patterson, J.W., Cyanide, in Industrial Wastewater Treatment Technology, 2nd ed., ButterworthHeinemann, Boston, MA, 1985, p. 115. 6. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd, London, 1991. 7. Mudder, T., Cyanide Monograph, Mining Journal Books, Ltd, London, 1998. 8. Cohn, J.G., Stern, E.W., and Etris, S.F., Gold and gold compounds, in Kirk–Othmer Encyclopedia of Chemical Technology, online edition [April 16, 2001], John Wiley & Sons, New York, 2001. 9. USDI, World gold: a minerals availability appraisal, SP 24-94, U.S. Department of Interior, Bureau of Mines, Washington, DC, 1994. 10. VonMichaels, H., Role of cyanide in gold and silver recovery, in Proceedings of the Conference on Cyanide and the Environment, Tucson, AZ, 1984, p. 51.

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11. Young, C.A., Cyanide: just the facts, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 97. 12. EA, Cyanide Management, Australian Department of the Environment and Heritage, http://www.deh. gov.au/industry/industry-performance/minerals/booklets/cyanide/index.html, accessed: March 31, 2005. 13. Mudder, T., Fox, F., Whitlock, J.L., and Marshall, S., Development of site-specific discharge criteria through toxicological testing, in Proceedings of the Conference on Cyanide and the Environment, Tuscon, AZ, 1984, p. 109. 14. Beebe, R.R., Process considerations before and after failure of the Omai Tailings Dam August 10 to 24 1995, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 3. 15. Bigelow, R.C., Plumlee, G.S., and Edelman, P., The Summitville Mine and its downstream effects, Open File Report 95-23, U.S. Geological Survey, Denver, CO, http://pubs.usgs.gov/of/1995/ ofr-95-23/summit.htm, 1995. 16. Egan, T., The death of a river looms over choice for Interior post, New York Times, January 7, 2001. 17. WSN, Group support new bill banning cyanide use in Wisconsin mines, Wisconsin Stewardship Network, http://www.nocrandonmine.com/wsn/mining/cyanidebannews.html, 2004. 18. Hackl, R.P. and Wright, F.R., Bioleaching of refractory gold ores — out of the lab and into the plant, in Proceedings of the Biohydrometallurgy Symposium, Jackson Hole, WY, 1989. 19. Han, K.N. and Fuerstenau, M.C., Ammonia leaching of gold and gold-bearing ores, in Proceedings of the XX Intern. Mineral Processing Congress, 1997. 20. Hiskey, J.B. and Atluri, V.P., Dissolution chemistry of gold and silver in different lixiviants, Mineral Process. Extractive Metall. Rev., 4, 95, 1988. 21. Hunter, R.M., Stewart, F.M., Darsow, M., Fogelsong, M., Mogk, D., Abbott, E., and Young, C.A., A new alternative to cyanidation: biocatalyzed bisulfide leaching, in Proceedings of the 3rd International Conference on Minerals Bioprocessing in Mining, Big Sky, MT, 1996. 22. Veto, R.H. and McNulty, T.P., Evaluation of non-cyanide technologies for processing ore of the McDonald gold deposit, Montana, in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 83. 23. Wan, R.Y., Importance of solution chemistry for thiosulfate leaching of gold, in Proceedings of the 1997 World Gold Conference, Singapore, Australian Inst. of Mining and Metallurgy, 1997, p. 159. 24. Anderson, C.G. and Nordwick, S.M., Pretreatment using alkaline sulfide leaching and nitrogen species catalyzed pressure oxidation on a refractory gold concentrate, in Proceedings of the EPD Congress, The Minerals, Metals and Materials Society, Warrendale, PA, 1996. 25. Ritchie, I.M., Nicol, M.J., and Staunton, W.P., Are there realistic alternative to cyanide as a lixiviant for gold at the present time? in Cyanide: Social Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 427. 26. Nicol, M.J. and O’Malley, P., Recovery of gold from thiosulfate solutions and pulps with ion-exchange resins, in Cyanide: Social, Industrial and Economic Aspects, Young, C.A., Twidwell, L.G., and Anderson, C.G., Eds., The Minerals, Metals and Materials Society, Warrendale, PA, 2001, p. 469. 27. Mudder, T., Editorial Comment: Minerva, Mining Environ. Manage., 9, 3, 2001. 28. Mudder, T. and Goldstone, A., Recovery of cyanide from slurries, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Book, Ltd, London, 1998. 29. Stevenson, J., Botz, M., Mudder, T., Wilder, A., Richins, R., and Burdett, B., Recovery of cyanide from mill tailings, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 30. CSU, Proceedings of the Conference on Cyanide and the Environment, Colorado State University, http://www.unr.edu/mines/mlc/conf_workshops/cyanide.html, accessed: March 25, 2005. 31. Dubey, S.K. and Holmes, D.S., Biological cyanide destruction mediated by microorganisms, World J. Microbiol. Biotechnol., 11, 257, 1995. 32. Bernoth, L., Firth, I., McAllister, P., and Rhodes, S., Biotechnologies for remediation and pollution control in the mining industry, Miner. Metall. Proc., 17, 105, 2000.

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33. Schlesinger, M., Electroplating, in Kirk–Othmer Encyclopedia of Chemical Technology, online edition [June 18, 2004], John Wiley & Sons, New York, 2004. 34. USEPA, A guide to cleaner technologies and alternative metal finishes, EPA-625/R-94-007, U.S. Environmental Protection Agency, Washington, DC, 1994. 35. Abrams, F., Land disposal restrictions: advance notice of proposed rulemaking, EPA Docket Number F-2000-LRRP-FFFFF, comments from IPC-Association of Connecting Electronic Industries, to U.S. Environmental Protection Agency, RCRA Information Center, IPC-ACEI, http://www.ipc.org/3.0_Industry/3.4_EHS/LDRcomments.doc, accessed: April 3, 2005. 36. USEPA, 180-day accumulation time under RCRA for waste water treatment sludges from the metal finishing industry: final rule (40 CFR Part 262), U.S. Environmental Protection Agency, Washington, DC, Fed. Regist., 65, 12378, 2000. 37. Stevens, W.F., The complete oxidation of cyanide from metal finishing operations, in Proceedings of the Conference on Cyanide the Environment, Tucson, AZ, 1984, p. 501. 38. Zabban, W. and Helwick, R., Cyanide waste treatment technology — the old, the new and the practical, Plating Surf. Finish, 67, 56, 1980. 39. Robine, R. and Lenglen, M., The Cyanide Industry, John Wiley & Sons, New York, 1906. 40. USS, The Making, Shaping and Treating of Steel, 9th ed., McGannon, H.E., Ed., U.S. Steel Corp., Pittsburgh, PA, 1971. 41. Wong-Chong, G.M. and Caruso, S.C., An evaluation of the treatment and control technology recommended for blast furnace (iron) wastewater, Carnegie Mellon Research Institute report to American Institute of Iron and Steel, Pittsburgh, PA, 1976. 42. Sawyer, B., Zenz, D.R., Lue-Hing, C., Lordi, D.T., and Hill, R., Realistic limits for water toxics: greater chicago demonstrates need for site-specific standards, Water Environ. Technol., 10, 57, 1998. 43. Wilson, P.J. and Wells, J.H., Coal, Coke and Coal Chemicals, McGraw-Hill, New York, 1950. 44. Ganczarczyk, J.J., State-of-the-art in coke plant effluent treatment, CRC Crit. Rev. Environ. Control., 13, 103, 1983. 45. Luthy, R.G., Treatment of coal coking and coal gasification wastewaters, J. Water Pollut. Control Fed., 53, 325, 1981. 46. Luthy, R.G. and Jones, L.D., Biological oxidation of coke plant effluent, J. Environ. Eng. Div., ASCE, 106, 847, 1980. 47. Wong-Chong, G.M., Caruso, S.C., and Patalis, T.G., Treatment and control technology for coke and plant wastewaters, in Proceedings of the AIChE 84th National Meeting, Atlanta, GA, 1978. 48. Wong-Chong, G.M., Biological degradation of cyanide in complex industrial wastewaters, in Proceedings of the International Symposium on Biohydrometallurgy, BIOMINET, CANMET Mining and Mineral Sciences Laboratories, Natural Resources Canada, Jackson Hole, WY, 1989. 49. Luthy, R.G., Bruce, S.G., Walters, R.W., and Nakles, D.V., Cyanide and thiocyante in coal gasification wastewaters, J. Water Poll. Control Fed., 51, 2267, 1979. 50. Wong-Chong, G.M. and Caruso, S.C., Advanced biological oxidation of coke plant wastewaters for the removal of nitrogen compounds, Carnegie Mellon Research Institute report to the American Iron and Steel Institute, Pittsburgh, PA, 1977. 51. Anthonisen, A.C., Loehr, R.C., Prakasam, B.S., and Srinath, E.G., Inhibition of nitrification by ammonia and nitrous acid, J. Water Poll. Control Fed., 48, 835, 1976. 52. Hockenberry, M.R. and Grady, C.P.L., Inhibition of nitrification — effects of selected organic compounds, J. Water Poll. Control Fed., 49, 768, 1977. 53. Sharma, B. and Ahlert, R.C., Nitrification and nitrogen removal, Water Res., 11, 897, 1977. 54. Ghosh, R.S., Dzombak, D.A., and Luthy, R.G., Equilibrium precipitation and dissolution of iron cyanide solids in water, Environ. Eng. Sci., 16, 293, 1999. 55. Happold, F.C., Isolation and characterization of an organism oxidizing thiocyanate, J. Gen. Microbiol., 10, 261, 1954. 56. Happold, F.C. and Key, A., The bacterial purification of gas works liquors. II. The biological oxidation of ammonium thiocyanate, Biochem. J., 31, 1323, 1937. 57. Beychock, M.R., Aqueous Wastes from Petroleum and Petrochemical Plants, John Wiley & Sons, London, 1976. 58. Huff, J.E. and Bigger, J.M., Cyanide removal from refinery wastewater using powdered activated carbon, EPA-600/2-80-125, U.S. Environmental Protection Agency, Washington, DC, 1980.

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59. Speight, J., Petroleum refinery processes, in Kirk–Othmer Encyclopedia of Chemical Technology, online edition [Dec. 4, 2000], John Wiley & Sons, New York, 2000. 60. Brix, K.V., Cardwell, R.D., Henderson, D.G., and Marsden, A.R., Site-specific marine water-quality criterion for cyanide, Environ. Toxicol. Chem., 19, 2323, 2000. 61. Hansen, Studies in support of alternate cyanide effluent limits for four San Francisco Bay area refineries, Report by S.R. Hansen and Associates, The Joint Refinery Cyanide Study Group: Tosco Avon, Shell Martinez, Exxon Benicia, and Unocal San Francisco Refineries, Berkeley, CA, 1990. 62. Wulf, R., Resources Refinery, Coffeyville, KS, personal communication, 2005. 63. API, Sour water stripping evaluation Publication 927, Report WBWC 3064, American Petroleum Institute, Washington, DC, 1972. 64. Urban, D., Frisbie, S., and Croce, S., Compliance strategy for cyanides in petroleum refinery wastewater: I. Source characterization and treatment, Environ. Prog., 16, 171, 1997. 65. Baker, D.C. and Chou, C.C., Cyanide occurrence and treatment in the petrochemical industry, in Proceedings of the Conference on Cyanide and the Environment, Tucson, AZ, Colorado State University, Fort Collins, CO, 1984, p. 379. 66. Wallschlager, D. and Bloom, N.S., Determination of selenite, selenate and selenocyante in waters by ion chromatography — hydride generation — atomic fluorescence spectrometry (IC-HG-AFS), J. Anal. Atom. Spectrom., 16, 1322, 2001. 67. Manceau, A. and Gallup, D.L., Removal of selenocyanate in water by precipitation: characterization of copper–selenium precipitate by x-ray diffraction, infrared and x-ray absorption spectroscopy, Environ. Sci. Technol., 31, 968, 1997. 68. Meng, X.G., Bang, S., and Korfiatis, G., Removal of selenocyanate from water using elemental iron, Water Res., 36, 3867, 2002. 69. Twidwell, L., McCloskey, J., Miranda, P., and Gale, M., Potential technologies for removing selenium from process and mine wastewater, in Proceedings of the REWAS-99, San Sebastian, Spain, 1999. 70. Prince, R., ExxonMobil, personal communication, 2005. 71. Whitlock, J.L. and Mudder, T., The Homestake wastewater treatment process: biological removal of toxic parameters from cyanidation wastewaters and bioassay effluent evaluation, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 72. Wedl, D.J. and DFaulk, R.J., Cyanide destruction in plating sludges by hot alkaline chlorination, Metal Finish., 89, 33, 1991. 73. Wunsch, A. and Nagel, R., Cyanide destruction with ozone and per-acids: special problems in particular cases, Galvanotechnik, 82, 3552, 1991. 74. Stolyarova, V.E., Yanbukhtina, R.A., Laskin, B.M., Povelikina, L.N., Anukhina, I.A., and Tsareva, O.A., Treatment of concentrated wastes containing cyanides and chromium(III) compounds, Russian J. Applied Chem., 73, 89, 2000. 75. Depoisoning of cyanide and chemical nickel, Oberflachen Werfkstoffe/Sufaces Materiaux, 1997, p. 13. 76. Baldwin, P.C., Electrolytic plate-out of scrapped cyanide copper baths, Met. Finish, 91, 13, 1993. 77. Szpyrkowicz, L., Zilio-Grandi, F., Kaul, S.N., and Polcaro, A.M., Copper electrodeposition and oxidation of complex cyanide from wastewater in an electrochemical reactor with a Ti/Pt anode., Indus. Eng. Chem., 39, 2132, 2000. 78. Hofseth, C.S. and Chapman, T.W., Electrochemical destruction of dilute cyanide by copper-catalyzed oxidation in a flow-through porous electrode, J. Electrochem. Soc., 146, 1999, 1999. 79. Blatt, W. and Schneider, L., Demetallization and cyanide oxidation, Metalloberflache, 50, 886, 1996. 80. Monser, L. and Adbhoum, N., Modified activated carbon for removal of copper, zinc, chromium and cyanide from wastewater, Sep. Purif. Technol., 26, 137, 2002. 81. Sanchez, H., Chainet, E., Nguyen, B., Ozil, P., and Meas, Y., Electrochemical deposition of silver from low cyanide concentration bath, J. Electrochem. Soc., 143, 2799, 1996. 82. Chen, Y.S., You, C.G., and Ying, W.C., Cyanide destruction by catalytic oxidation, Metal Finish., 89, 68, 1991. 83. Slawski, K., Wedzicha, L., and Mromlinska, Z., Investigations of recovery of silver and cyanide from the effluents of silver plating process, with use of ion–silver electrodes, Rudy I Metale Niezelazne, 37, 248, 1992. 84. Wong-Chong, G.M. and Caruso, S.C., An evaluation of the treatment and control technology recommended for blast furnance (iron) wastewater, Carnegie Mellon Research Institute report to American Institute of Iron and Steel, Pittsburgh, PA, 1976.

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Management in 27 Cyanide Groundwater and Soil Rajat S. Ghosh, David V. Nakles, David A. Dzombak, and George M. Wong-Chong CONTENTS 27.1

Groundwater Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.1.1 Establish Basis and Assessment Methods for Compliance. . . . . . . . . . . . . . . . . . . . . . 27.1.2 Concentration and Distribution of Cyanide Species in Impacted Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.1.3 Management Decision Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.2 Management of Cyanide-Impacted Soil and Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.2.1 Establish Basis and Assessment Methods for Compliance. . . . . . . . . . . . . . . . . . . . . . 27.2.2 Concentration and Distribution of Cyanide Species in Impacted Soils/Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.2.3 Management Strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.3 Case Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.3.1 Manufactured Gas Plant Site Closure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.3.2 Spent Heap Leach Pad Closure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.3.3 Spent Potlining Leachate Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.3.4 Electroplating Site Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27.4 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

572 572 573 574 575 576 578 578 579 579 581 582 584 586 586

Cyanide has been observed in both soil and groundwater at sites associated with aluminum smelting, manufactured gas plants (MGP), electroplating, and the mining and processing of metal ores. Cyanide is one of the most common chemicals detected at Superfund sites [1]. As discussed throughout this book, the forms of cyanide that are present in these media vary from site to site. Sites with cyanide contamination typically have groundwater with detectable concentrations of free cyanide (HCN, CN− ), weak acid dissociable (WAD) metal–cyanide complexes, available cyanide, and strong-acid-dissociable metal–cyanide complexes. However, the relative proportions of these various species of cyanide vary based on the industry-specific waste type. For example, cyanide in groundwater at MGP and aluminum production sites is predominantly iron–cyanide complexes, with relatively small fractions of WAD cyanide and almost no free cyanide. In contrast, metal mining and electroplating sites usually contain larger proportions of the WAD and free cyanide in addition to the complex metal (other than iron) cyanide compounds. The management of cyanide-impacted soil and groundwater is governed by specific regulatory criteria, such as soil quality criteria, groundwater and surface water quality criteria, and other toxicological criteria (Chapter 18). In the United States, many states have criteria for groundwater and soil quality, which often are linked to drinking water maximum contaminant levels. Noncompliance with these stated criteria results in violations, which usually requires mitigation by active treatment. 571

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In situations where there are no specific criteria or guidelines, which is often the case for cyanide-impacted soils, it is necessary to use standard risk assessment techniques as discussed in Chapters 16 and 17. These techniques consider the form and concentration of the cyanide that is present, the various routes of contact that can lead to exposure of the receptor to the cyanide-impacted media, and the effects that this exposure has on the receptor. For human health risk assessment, the routes of exposure that are usually of concern are direct dermal contact, ingestion, and inhalation. This chapter considers the management options that are available for cyanide-impacted groundwater and soil. It emphasizes the need for proper characterization of the impacted media and the use of that information for the assessment of compliance with regulations or the determination of risk to public health and the environment. Several case studies are presented to illustrate the usefulness of some of the management options.

27.1 GROUNDWATER MANAGEMENT Management of cyanide-impacted groundwater requires a tiered approach. The first step involves the establishment of a basis for compliance and the analytical methods for determining it. The second step involves a detailed understanding of the concentration and distribution of various cyanide forms at the site through careful choice of species-specific analytical methods discussed in Chapter 7. These steps will then lead to a management decision framework, the outcome for which can either be treatment of cyanide-impacted waters or risk-based management via a combination of factors that may include the following: (i) application of appropriate analytical method to determine compliance, (ii) institutional controls, and (iii) fate and transport modeling to determine the risk at point of compliance by considering various exposure pathways. The rest of this section presents a detailed discussion of this tiered approach.

27.1.1 ESTABLISH BASIS AND ASSESSMENT METHODS FOR COMPLIANCE The first step in the management of cyanide-impacted groundwater is to establish the basis for compliance and the analytical methods for determining it. Most regulatory criteria for cyanide in water are based on the concentration of free cyanide that is present, though monitoring requirements often specify total cyanide or other forms (see Chapter 18). This is because free cyanide (HCN, CN− ) is the toxicological species of primary concern (Chapters 13 to 17). However, the regulations specify the use of United States Environmental Protection Agency (USEPA)-approved analytical methods to assess compliance, that is, total cyanide [2], available cyanide [3], or cyanide amenable to chlorination (CATC) [4]. Consequently, the first step in managing cyanide-impacted groundwater is to establish with the regulatory agencies that free cyanide is the cyanide species of interest and should form the basis for compliance assessments. Next, it is important to establish the analytical method or methods that will be used to measure the concentration of free cyanide. It is certainly appropriate to use total cyanide analyses for screening assessments, since this represents a conservative approach in that total cyanide captures metal–cyanide species as well as free cyanide. Stated differently, if compliance with the regulations is achieved assuming that all of the cyanide measured using total cyanide analysis is free cyanide, then compliance based on the actual concentration of free cyanide is assured. If total cyanide concentrations exceed the regulatory criteria, then analytical methods should be proposed that provide the best estimate of the concentration of free cyanide. As of this writing, the only USEPA-approved methods that can be used for measurement of free cyanide are the CATC [4] and OIA-1677 [3] procedures. Both of these methods include cyanide compounds other than free cyanide, but they are less conservative estimates (i.e., provide better estimates of the free cyanide concentration) than that provided by the total cyanide procedure. The

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former method (CATC) is promulgated under both Clean Water Act (CWA) and Safe Drinking Water Act (SDWA) and can be used for applications involving all waters, that is, surface water, groundwater, or drinking water. The latter method (OIA-1677) currently has more limited applications and is promulgated only under CWA and therefore cannot be used to assess compliance with drinking water standards. However, of these two methods, OIA-1677 is obviously the preferred method. It is more reliable, with much better reproducibility, precision, and accuracy (USEPA, 1999). The American Society for Testing and Materials (ASTM) analysis for WAD cyanide [5] or the equivalent Standard Methods technique [6] will yield results that are very close to those of OIA-1677. Even though WAD cyanide analysis is not an USEPA-approved method, its use is preferable over CATC and should be suggested to any agency as a surrogate for the OIA-1677 if they are inclined to use the CATC method because of their familiarity with it [7]. A simple, reliable, and accurate method to use for compliance measurements of free cyanide is the ASTM microdiffusion method [8]. This method measures predominantly the concentration of free cyanide but can also capture some metal–cyanide species that have low stability and readily dissociate under the weak acid test conditions [7]. This method has been validated through an interlaboratory study and is being evaluated for approval by the USEPA. A three-tiered compliance assessment built around the existing analytical methods and their approval status will be useful in many situations. This involves first-tier screening using total cyanide analyses, followed by a second-tier screening using available cyanide [3] or WAD cyanide analyses, and then, if necessary, a third-tier using the microdiffusion analysis. When the microdiffusion method is approved for regulatory use, it could be used in the second step in place of the available cyanide and WAD cyanide measurements. Most of the water-based regulatory criteria for cyanide are in the form of specific groundwater quality criteria, groundwater/surface water interface (GSI) criteria, or surface water quality criteria. The GSI criteria are meant to protect surface water quality in those instances where there is a direct connection between the groundwater and surface water. When surface water discharge becomes a part of the compliance equation, there is another dimension to the compliance assessment beyond a direct comparison of the concentrations of the cyanide species to the regulatory criteria. This involves the potential transformation of the complex metal cyanides to free cyanide via photodecomposition (see Chapter 9). Many states deal with the issue of surface discharge by making the conservative assumption that all of the metal–cyanide complexes will photodissociate to form free cyanide. For example, the State of Michigan has instituted a GSI criterion of 5.2 ppb (as available cyanide using USEPA Method OIA-1677) [9], which is the USEPA chronic fresh water quality criterion. However, based on information provided in Chapters 5 and 9, it is evident that while dissociation can be rapid under particular solution and irradiation conditions, it depends on various physicochemical factors and could be significantly impeded in the field [10]. Moreover, there is evidence that the free cyanide that is formed from the photodissociation biodegrades more rapidly than it volatilizes ([11,12]; and Chapter 9). Currently, there are only limited field data available regarding the fate of cyanide in surface waters. However, a fate and transport model has been developed by researchers at Clarkson University that incorporates the available data and can predict the fate of free cyanide in surface water following discharge of a complex mixture of cyanide compounds (Chapter 9). In summary, where groundwater discharges into surface water, compliance with surface water criteria can be determined through direct measurements of free cyanide concentrations in the receiving water or by using groundwater quality and a cyanide fate and transport model.

27.1.2 CONCENTRATION AND DISTRIBUTION OF CYANIDE SPECIES IN IMPACTED GROUNDWATER The concentration and distribution of cyanide species in groundwater will vary with source type and site conditions. Groundwaters impacted by leaching of cyanide-containing source materials from

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TABLE 27.1 Typical Concentration and Distribution of Cyanide Species in Groundwater at Industrial Sites

Industry Manufactured gas plant Aluminum production Mining Metal finishing

Total cyanide concentration range, mg/l

Strongly complexed, %

Weakly complexed, %

Free cyanide, %

0.2 – 20 0.2–50 0–30 0–5

>90 >90 >60 60–80

<8 <8 <40 60–80

0–2 <1 <5 20–40

Source: Information from Kjeldsen, P., Water, Air, Soil Pollut., 115, 279, 1999; Ghosh, R.S., Nakles, D.V., Murarka, I., and Neuhauser, E.F., Environ. Eng. Sci., 21, 752, 2004; Blayden, L.C., Hohman, S.C., and Robuck, S.J., Proceedings of Light Metals 1987, The Minerals, Metals and Material Society, Warrandale, PA, 1987; Kimmerle, F.W., Girard, P.W., Roussel, R., and Tellier, J.G., Proceedings of Light Metals 1989, The Minerals, Metals and Material Society, Warrandale, PA, 1989; Smith A. and Mudder, I., The Chemistry and Treatment of Cyanidation Wates, Mining Journal Books, Ltd., London, 1991; Irwin, R.J., Environmental Contaminants Encyclopedia, National Park Service, Water Resources Division, Water Operations Branch, Fort Collins, CO, 1997.

MGP sites and aluminum smelting facilities contain cyanide that is strongly complexed with iron, with less than 10% available as free or WAD complexes [13–17]. In contrast, a significant portion of cyanide in mine tailings impacted groundwater is comprised of weak metal–cyanide complexes (20–40%), with free cyanide constituting less than 5% and the rest being strong cyanide complexes with iron and cobalt [18]. In groundwaters impacted with plating wastewaters and sludges, metalcomplexed cyanides could constitute 60–80% of the total cyanide with the remaining 20–40% being free cyanide [19]. Typical concentrations and speciation of cyanide in groundwater at these industrial sites are provided in Table 27.1. As shown in Table 27.1, metal-complexed cyanides dominate cyanide speciation in groundwater at almost all industrial sites; this is due to the reactivity of cyanide ion and the abundance of metals, especially iron, in soil. In addition, the fact that free cyanide is readily biodegradable in shallow soil-groundwater systems in comparison to any other complexed metal cyanides prevents its buildup in groundwater over time.

27.1.3 MANAGEMENT DECISION STRATEGY Following the establishment of the compliance criteria, selection of appropriate analytical methodology and performance of detailed site characterization, the next phase is to conduct a cyanide species-specific comparison with various compliance criteria to evaluate exceedances. For example, if compliance is determined on the basis of free cyanide concentrations, then diffusive cyanide determined by microdiffusion method [8] can be compared with the compliance criteria. Otherwise, more conservative analytical methods like available cyanide [3] or total cyanide by distillation can be used. Both temporal and spatial profiles for exceedances are taken into account while determining a suitable strategy for management. For situations where the groundwater–surface water connection is absent and only total cyanide exceeds the compliance criteria, a risk-based management strategy can be adopted that might include comprehensive monitoring and institutional controls, such as prohibition of land and water use within the plume area by putting up fences. However, if a groundwater–surface water pathway exists and there is potential for surface water photodissociation of complexed cyanides, then surface water quality modeling results could be used in conjunction

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with the cyanide analytical data to assess surface water quality impacts at one or more selected points of compliance. The last option available in managing cyanide impacts in groundwater is active treatment of either the source area or contaminated plume or both. This is particularly important when free cyanide concentrations exceed the compliance criteria. Often times, performing active remediation might prove to be economical, particularly in cases where there is a small release, comprised mostly of weak metal–cyanide complexes and the contaminated plume has not extended far from the source area or is very slow moving. Cyanide removal technologies for groundwater mostly involve ex situ pump and treat processes [20–22]. Above-ground treatment of groundwater bearing free cyanide is usually accomplished by alkaline chlorination [22], but for groundwaters with strongly complexed cyanide, usually in the form of iron–cyanide complexes, other approaches such as adsorption, precipitation, or UV/oxidation are needed. Treatment options are described in Chapters 19 to 24. In situ treatment approaches have also been implemented for cyanide in the subsurface. Kastman and Zimmerman [23] employed in situ alkaline chlorination at a sodium cyanide disposal site. More recently, the use of an in situ iron–sand permeable reactive barrier [24,25] as well as phytoremediation [26,27] have been investigated as innovative methodologies for in situ treatment of subsurface cyanide contamination. In considering treatment alternatives it is important to recognize that the extent of treatment achievable is dependent upon the cyanide species that are present. Treatment alternatives for the free, WAD, and available cyanide are generally more cost effective than those for the strong metal–cyanide complexes. This has important implications in that, as previously noted, the target analyte of concern should be the free cyanide. Since the free, WAD, and available cyanide concentrations represent better estimates of the concentration of free cyanide than do the concentrations of total and CATC, it is preferable to use the former methods to evaluate the effectiveness of any treatment process. For example, it is known that free cyanide is amenable to aerobic biological degradation while the strong metal–cyanide complexes are not. If the effectiveness of this treatment technology were gauged by its ability to treat total cyanide in a situation where the strong metal–cyanide complexes represented a significant portion of the total cyanide, it could be viewed as ineffective. On the other hand, if the performance evaluation focused only on the removal of free cyanide, it would most probably be considered acceptable. Once again, this highlights the importance of establishing with the regulators the cyanide species that will form the basis for compliance and then using the appropriate analytical methods to make that assessment. In many instances, if this is properly done at the beginning of a project, it may be concluded that the water of concern is already in compliance and that no treatment is required.

27.2 MANAGEMENT OF CYANIDE-IMPACTED SOIL AND RESIDUALS Like groundwater, management of cyanide-impacted soils and residuals also requires a tiered approach. The first step involves the establishment of a basis for compliance using site-specific exposure pathway analysis and analytical methods for its determination. The second step involves a detailed understanding of the cyanide concentration and species distribution in the impacted soil/residuals through selection of appropriate extraction tests and analytical methods as discussed in Chapter 8. These steps will then lead to a management decision framework, the outcome for which can either be excavation and treatment of the cyanide impacted soils/residuals, or risk-based management via a combination of factors that may include the following: (i) application of appropriate extraction protocols and analytical methods to determine cyanide speciation and distribution in the matrix of interest; (ii) detailed cyanide characterization coupled with exposure pathway analysis to determine compliance; and (iii) institutional controls. This section presents a detailed treatment of the two-tiered approach.

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27.2.1 ESTABLISH BASIS AND ASSESSMENT METHODS FOR COMPLIANCE Existing criteria for cyanide in soils or residuals are based on the consideration of human health risk and three basic pathways of exposure. First, there is the potential for exposure by direct contact with the cyanide-impacted material, which can occur via direct dermal contact or incidental ingestion. Second, there is the protection of groundwater or surface water potentially impacted by leachate. In this case, the cyanide concentration in the soil/residual is regulated to ensure that cyanide does not leach from the solid matrix and produce concentrations in groundwater or surface water that exceed applicable criteria for these media. Finally, there is the protection with respect to the inhalation pathway. Each of the exposure pathways should be evaluated to ensure that the most appropriate leach methods and analytical procedures are applied for the soil or solid residual for the particular compliance assessment. Similar to the situation for groundwater, the assessment of compliance with these criteria require an understanding of the cyanide species that will be regulated and the analytical methods that will be used to make the measurements. However, when solid matrices are involved, it is necessary to address two additional steps in the analytical procedure, that is, sampling of the solid matrix and the extraction of cyanide from the solid matrix. It is very important to secure a representative sample for analysis. Because of the issue of heterogeneity in soil matrices, it is important to develop a soil sampling program that includes the necessary plans for acquisition of representative samples from a particular matrix. Soil sampling protocols need to be developed on a site-by-site basis. Guidance for soil sampling has been developed by the USEPA [28] and other regulatory agencies. There are some special considerations for mining sites [29]. Additional information is provided in Chapter 8. There are several methods that can be used to extract cyanide from a solid matrix for analysis, as described in Chapter 8, and for leaching solid samples for exposure and risk assessment purposes. The analytical extraction method or leaching method selected should reflect the environmental conditions associated with the specific exposure pathway of interest. Of particular importance are the pH and oxidation–reduction conditions of the extraction procedure. For example, the leaching that takes place in the human stomach occurs at a pH of about 2. Consequently, the cyanide exposure that results from ingestion should be assessed using an acidic leaching medium at or near pH 2. Selection of an appropriate extraction or leaching protocol requires careful consideration of the relevant environmental conditions and the information needed for site management or compliance decisions [30]. Compliance with direct contact criteria for cyanide-impacted soils and residuals is assessed by leaching the cyanide from the solid matrices, analyzing the leachate for the cyanide species of interest, and then determining the concentration of that cyanide species in the solid matrix, that is, mass of cyanide species in the leachate divided by the mass of soil from which it was leached. For example, the State of Michigan has direct contact criteria for cyanide of 12 mg/kg (as available cyanide using USEPA OIA-1677) for residential exposure scenarios and 250 mg/kg (as available cyanide using USEPA OIA-1677) for industrial exposure scenarios [31]. To determine compliance with these criteria following direct dermal contact, it may be appropriate to use a neutral leach method since the pH of the human skin is at a near-neutral pH condition, that is, pH of 5 to 6. Then, as previously stated, this leachate could be analyzed for free cyanide to produce a concentration in the solid that can be compared to specific human health, direct contact criteria. The State of Michigan recommends using a neutral leach protocol [32] to simulate direct contact scenarios for cyanide [33]. For ingestion, it would be more appropriate to use an acid leach (pH of 1.5 to 2), possibly followed by a slightly alkaline leach (pH of 7) to simulate the conditions of the stomach and intestines, respectively. These leachates would then be characterized in a manner identical to what was described above for direct dermal contact. It is interesting to note that there is a cyanide analytical method

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in the State of Massachusetts referred to as the physiologically available cyanide, or PAC [34]. This method was specifically designed to mimic the physiological conditions that are found in the human stomach, in particular, the pH which is fixed at 1.5. However, the temperature that is prescribed for this method is significantly higher than that of the human body. In spite of this limitation, the method does provide a better estimate of exposure to cyanide following human ingestion when compared to analysis/characterization based on total cyanide. A number of leaching tests have been developed by regulatory agencies for particular disposal scenarios, such as the toxicity characteristic leaching procedure (TCLP), which is designed to simulate conditions in a municipal waste landfill, and the synthetic precipitation leaching procedure (SPLP) in which leaching by acidic rainfall is simulated [35,36]. There are other tests, as discussed in Chapter 8 and elsewhere [29,30]. As emphasized by Kosson et al. [30], no single leaching test is appropriate for all situations. Leaching procedures must be selected with careful consideration of the site conditions and the goals of the assessment. As far as the analytical method is concerned, the most appropriate one should be used for determining the concentration of free cyanide in the leachate. The microdiffusion method [8] provides the best estimate of free cyanide, followed by the WAD [5] and available cyanide [3] methods. The analytical methods for CATC [4] and total cyanide [37] would be the least desirable methods for this purpose. Groundwater quality criteria, to which soil quality criteria are linked, are most stringent when the groundwater is discharging into surface water because of the extremely low chronic and acute cyanide criteria for ecological receptors, that is, 5.2 and 22 µg/l as free cyanide (see Chapters 14 and 18). The next most stringent criteria are for the protection of human health, addressing groundwater that may be ingested by humans or groundwater with which humans may come into direct contact, for example, dermal exposure during excavation operations. Soil quality criteria have been developed to provide human health protection for exposure to groundwater that has contacted the soil. Soil criteria for cyanide developed by the State of Michigan illustrate the consideration of different possible exposure pathways. The Michigan soil cyanide criteria are: 0.1 mg/kg (as available cyanide using USEPA OIA-1677) for protection of the GSI (i.e., instances where groundwater discharges into surface water); 4 mg/kg for protection of drinking water (as available cyanide using USEPA OIA-1677); and 250 mg/kg (as available cyanide using OIA-1677) for protection of groundwater in direct contact with humans [38]. The last pathway of concern is the inhalation pathway. This pathway merits consideration in risk assessments for cyanide in soil and groundwater, as hydrogen cyanide if present or generated in the waste can volatilize and is very toxic. The generation of HCN gas from a solid matrix involves two steps: (1) the cyanide must be released from the solid to the aqueous phase, and (2) the dissolved species must be decomposed to form free cyanide, which is in the form of HCN under neutral to acidic conditions and subject to volatilization. Because of the multiple physical–chemical steps and the rate-limited mass transfer processes involved, not all cyanide-containing residuals or wastes will yield HCN gas when subjected to mid-range or acidic pH conditions. For example, cyanide-impacted solids from MGP or aluminum smelting sites are usually dominated by iron–cyanide compounds that have very low solubility under acidic conditions and therefore will not leach to an appreciable extent under these conditions. On the other hand, sludges or solid residuals from other industries, such as the metal plating industry, might release free cyanide under these conditions because of the dominance of weak metal–cyanide compounds and dissolved complexes. In general, the generation of HCN gas from cyanide-impacted solids or sludge is unlikely under typical environmental conditions, but needs to be investigated. A quick assessment of the potential for this to occur can be made based upon the cyanide species that are present in the solid. If strong metal– cyanide complexes, such as iron–cyanide species, are the dominant species released to solution from the solid, dissociation will be very slow in the dark for all pH conditions and HCN gas production will be negligible. On the other hand, if weakly complexed metal cyanides are dominant, the solid should probably be investigated further.

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27.2.2 CONCENTRATION AND DISTRIBUTION OF CYANIDE SPECIES IN IMPACTED SOILS/RESIDUALS The concentration and distribution of cyanide species in soils and residuals will depend on the nature of the extraction conditions. Theis et al. [15] analyzed alkaline extracts of spent oxide box wastes at MGP sites and found ferrocyanide complexes to dominate the cyanide speciation in the extracts (>80%). Other cyanide species of interest included thiocyanate (SCN− ) and trace amounts of free cyanide. A similar study was also performed by Ghosh et al. [17], who subjected soils from ten different MGP sites in the State of New York to different leaching conditions and then performed detailed cyanide speciation on the leachates. Each site soil was subject to acid (pH = 1.5), deionized (DI) water (pH = 5.5), and caustic (pH = 14) extraction conditions and the resulting leachates were analyzed for free cyanide, total cyanide, available cyanide, and individual metal–cyanide complexes. The results of their analysis indicated that under acid leaching conditions, an iron– pentacyano complex [Fe(CN)5 NHCH3 ]4− , a strong acid-dissociable cyanide, dominated the leachate speciation at all the sites. However, under caustic leaching conditions, ferrocyanide complex was the predominant form of cyanide, while both the iron–cyanide complexes dominated speciation under the DI water leach condition. Leachates from aluminum spent potlining material under alkaline conditions are also dominated by ferrocyanide complexes with less than 10% available as free or WAD complexes [13,14]. The commonly used regulatory leaching tests such as the TCLP and SPLP tests have been employed, along with other leaching tests, to evaluate the leaching characteristics of solid wastes from metal mining operations [29]. Much of this testing has been directed at evaluation of metals leaching (e.g., [39,40]). There is very little information available in the literature on the cyanide leaching characteristics of mining wastes under environmental (i.e., nonproduction process) conditions. The limited literature available on application of leach tests to mining wastes suggests, as others have noted [41], that there has been a relatively little development of leach testing focused on metal mining wastes for regulatory assessments in the United States. This perhaps is related to the exclusion of mining wastes from the hazardous waste regulations under Resource Conservation and Recovery Act (RCRA). The mining industry did challenge, unsuccessfully, the application of the TCLP test to mineral processing wastes [42].

27.2.3 MANAGEMENT STRATEGY Following the establishment of a basis for compliance using site-specific exposure pathway analysis and selection of appropriate solid extraction protocol, the next step is to apply the appropriate analytical methodology to measure the target species for comparison with established compliance criteria to evaluate exceedances. For example, if ingestion is considered the primary pathway of exposure, then an acid leach (mimicking the GI tract conditions) or a more conservative DI water leach protocol may be used to extract the solid. For direct dermal exposure, a neutral pH leach could also be applied. For example, the State of Michigan recommends the use of a neutral leach protocol [32] to simulate direct contact scenarios for cyanide. Following the generation of the extract, if compliance is determined on the basis of free cyanide concentrations, then measurements by applying the microdiffusion method [8] to the soil extract can be compared with the compliance criteria; otherwise more conservative methods, like USEPA Method OIA-1677 [3] or total cyanide by distillation [2] can be used. Failure to comply with these criteria most often requires management of cyanide-impacted soils/residuals via deed restriction or by excavation and subsequent disposal at hazardous waste landfill sites. The last option available in managing cyanide impacts in soils/residuals is however, pursuing active on-site treatment. Treatment of cyanide-impacted soils and solids can be achieved in one of two ways: (1) the solid matrix can be directly treated as a whole, or (2) the cyanide can be extracted from the solid after which the cyanide-laden extract can be subjected to treatment. If the extraction route is pursued, then

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the water treatment options outlined in Chapter 19 apply to the cyanide extracts. However, some consideration will have to be made for the conditions that were required to achieve an efficient extraction of the cyanide from the solid, for example, elevated pH. Following the generation of the extract, if treatment is aimed at destroying free cyanide and weak metal–cyanide complexes, less energyintensive technologies like alkaline chlorination, ozonation, or microbiological treatment may be applicable. On the other hand, if treatment is aimed at destroying strong metal–cyanide complexes, energy-intensive, destructive technologies like incineration, high temperature alkaline hydrolysis, or photocatalytic oxidation are required. If the solid is treated as a whole, the treatment options are more limited. For the most part, the only viable options are direct incineration or thermal desorption. Both processes convert the cyanide to carbon dioxide and nitrogen oxide, the former doing it in one step and the latter, in two steps. Other technologies, such as solidification or stabilization, have not been fully tested at this time and their effectiveness of treatment is not known. A summary of the treatment technologies and others that might be effective for managing cyanide-impacted solids are presented in Chapters 19 to 23.

27.3 CASE STUDIES 27.3.1 MANUFACTURED GAS PLANT SITE CLOSURE This case study provides a summary of the case history for an MGP site in the upper Midwest region of the United States [16,43]. Phases I and II environmental investigations conducted at this site in 1991 and 1992 indicated the presence of MGP tars in the gas holder area as well as cyanide impacts in the groundwater and soil. However, no cyanide-bearing source materials (e.g., oxide box residuals) were located. Based on these investigations, a soil remediation study was undertaken in 1993. As part of the excavation operations during that study, greenish–blue stained soil was discovered adjacent to the tar separator tanks and was found to contain cyanide compounds. This material was believed to be the source of cyanide to groundwater at the site, and was, therefore, removed during the study. In February 1995, the site remediation consultant, on behalf of the site owner, prepared two reports for the State regulatory agency. One report presented the results of the excavation and soil treatment activities that were conducted at the site in 1992 and 1993, and requested that unsaturated soils be closed out pursuant to the State regulatory requirements. The other report (Feasibility Study Report) outlined a course of action to address the presence of cyanide in groundwater remaining at the site following the soil remediation efforts. A two-phased approach was adopted to address the cyanide groundwater impacts. Phase I included the implementation of a groundwater monitoring program to determine if natural attenuation processes were capable of reducing cyanide concentrations in groundwater following the source removal activities. Phase II included the evaluation of additional remedial actions if natural attenuation could not sufficiently reduce cyanide concentrations in groundwater. In May 1995, the State regulatory agency approved the close out of the unsaturated soil portion of the site and requested of the site owner that the Phase I groundwater monitoring be conducted quarterly for two years, and that the program includes analyses for the following constituents: PAHs, total cyanide, CATC, and WAD cyanide. Around the same time, the site owner initiated a research project at the MGP site to investigate the fate and transport of MGP-derived cyanide in groundwater. The Phase I postremediation groundwater monitoring program conducted from September 1995 to September 1997 resulted in groundwater detections of cyanide throughout the two-year sampling period. However, no clear decreasing or increasing trend was identified. Based on the results of the Phase I program and the preliminary information being generated from the research, at the conclusion of the two-year monitoring program the site owner elected to extend the Phase I groundwater

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monitoring program until the research program was completed. This decision was made so that additional data could be collected to allow for a more appropriate evaluation of the postremediation groundwater conditions before making any recommendations for groundwater management options at the site. The cyanide analytical methods used in the program included total cyanide; WAD cyanide; free cyanide by microdiffusion; and available cyanide (i.e., USEPA Method OIA-1677). As part of the research program, a Geoprobe plume delineation investigation was conducted in November of 1995. Following this investigation, additional groundwater monitoring wells were installed that resulted in a total network of 41 permanent wells, which were sampled quarterly to provide an assessment of groundwater quality from July 1996 through April 2000. During this quarterly monitoring, groundwater from both onsite and offsite locations consistently contained detectable concentrations of cyanide. Results from this monitoring indicated that the cyanide in the groundwater was primarily in the form of iron–cyanide complexes (>96%), that these complexes are stable under the conditions of the aquifer, and they are transported as nonreactive solutes in the sand–gravel aquifer material [16]. Based on the cyanide concentration trends observed in groundwater at the site, two additional onsite areas were identified with the potential of having additional cyanide soil impacts. These areas were investigated with the completion of 27 soil borings in January, 2000. No cyanide-bearing oxide box residuals were found during this investigation and total cyanide was detected in only a few of the submitted samples, confirming that the large majority of the cyanide-bearing oxide box residuals at the site had been removed during the excavation in 1992 and 1993. However, a number of disperse stringers of residual reprecipitated cyanide were observed in the vadose zone and saturated zone. Although the majority of the source materials had been removed, these residual stringers continued to act as a dispersed source of cyanide to groundwater. Based on the findings of the additional investigation and the results of the five-year groundwater monitoring program, closure was requested for the groundwater and saturated soil portion of the MGP site pursuant to regulatory requirement. The justification for this request was as follows. The additional source characterization conducted at the site revealed that only dispersed, residual reprecipitated cyanide solids were present in the subsurface and that the majority of the source material had been removed. Due to the dispersed nature and low concentrations of these reprecipitated materials, it was determined to be impractical to achieve any additional benefits to groundwater quality through additional excavation at the site. Since the majority of the source was removed in the 1992 and 1993 excavations and only dispersed residuals remained, it was concluded that no additional removal actions were warranted at the site. Postremediation groundwater monitoring had been conducted during 20 separate events at the site over a 5-year period and had shown minimal cyanide impacts, though the source area continues to leach cyanide species from the residual dispersed material present. Only four wells (three onsite and one offsite) out of 41 ever demonstrated exceedances (based on WAD cyanide as a conservative estimate of free cyanide) of the 0.2 mg/l State groundwater standard during the 5-year monitoring period. The observed trends suggest that WAD cyanide concentrations at the site will, in the worst case, remain the same, or more likely to decrease as the residual reprecipitated materials are depleted over time. In conclusion, the extensive data that have been collected at the mid-western MGP site suggest that the majority of source materials have been removed and that only residual, dispersed cyanidebearing soils remain at the site. Furthermore, groundwater WAD concentrations remaining at the site are for the most part below the State groundwater standard and have remained steady over the last eight years, showing no offsite impacts above the State groundwater standard during the past several years of postremediation monitoring. Upon extensive review of the data provided by the site owner and their consultant to the State regulatory agency, the State granted conditional closure of the site in March, 2004. As part of the conditional closure, the State required the site owner to maintain certain monitoring wells in useable conditions for continued monitoring as part of an agreement between the City and the site owner.

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27.3.2 SPENT HEAP LEACH PAD CLOSURE Spent heap leach pad closure is required to meet both short- and long-term environmental objectives at a mining facility. Short-term objectives are directed toward treatment of all process solutions, stabilization of the heap solids, and erosion control. The long-term objective is to restore the site to its premined condition so that it can be used for livestock grazing, a wildlife habitat, or some other productive end use. In addition the closure must prevent any future environmental degradation [44,45]. The traditional approach to heap leach pad closure is to rinse the heaps with several pore volumes of water until the effluent rinse water meets required effluent standards [44,45]. The effluent from the initial rinse is often treated and recirculated with make-up water until the required effluent quality is achieved. Nutrients are sometimes added to stimulate microbial degradation of free and WAD cyanide [46]. However, this is generally only practicable and effective for smaller heap leach operations. The following challenges arise for large heaps: • Water requirements are substantial, and rinsing with multiple pore volumes may require a long time, even years. • The large volumes of contaminated rinse water generated require treatment. • Large heaps contain short-circuiting channels, which impact the effectiveness of the rinse. • While specified cyanide and pH limits in the rinse water may be met, other standards for metals and other inorganics may not be achieved [44]. An alternative to the traditional closure procedure for large heaps is the Biopass closure system [44,45,47]. The Biopass system is an alternative closure procedure to the traditional rinse approach described above. The purpose of the Biopass system is to provide a facility to degrade, stabilize, and contain contaminants contained in liquid seepage generated from the capped leach heap pad. The process promotes anaerobic biological degradation of WAD cyanide, with the production of hydrogen sulfide, and encapsulates the contained metal sulfides and other inorganic constituents of concern within the confines of the cell. In effect, there is no discharge from the closed heap leach unit. To this end, the treatment cell incorporates the following features, illustrated in Figure 27.1 and Figure 27.2: • A system of perforated plastic pipe along the bottom of the tailings pond associated with the heap leach pad to distribute seepage drainage from the capped leach heap. • A layer of organic matter (e.g., compost, wood chips, sawdust, and composted manure) on/around and above the distribution pipes. Native grasses and shrubs

Pressure release system Topsoil layer

4" Ø observation well

Reclaimed leach pad

Upper sand layer

1

3

From leach pad collection system Biopass system cell detail see Figure 27.2

Existing HDPE liner system

Lower sand layer Scale in feet 0

10

20

40

FIGURE 27.1 Schematic cross section diagram of the Biopass system. (Source: Mudder, T. et al., Cyanide Monograph, Mining Journal Books, Ltd, London, 1998. With permission.)

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Pressure relief sys (if required) 4924

Native grasses and shrubs Vegetated cover Recycled HDPE liner

4922

Elevation (ft)

4920

4918

Substrate layer

4916

4914 Geo textile 4912 Existing HDPE liner system 4910

FIGURE 27.2 Biopass system cell detail. (Source: Mudder, T. et al., Cyanide Monograph, Mining Journal Books, Ltd, London, 1998. With permission.)

• • • •

A porous geotextile fabric over the organic layer. 10 ft of spent heap leach fill material in the cell with a geotextile cover. A 1-ft layer of fill on a geogrid over the geotexile, with monitoring piping in this layer. Pressure relief pipes.

The depth of the organic layer is determined from an estimation of the organic requirements for anaerobic degradation of WAD cyanide and metals that are generated in the seepage from the closed heap leach pad [44,45,47].

27.3.3 SPENT POTLINING LEACHATE MANAGEMENT Spent potlining (SPL) is a by-product of the aluminum manufacturing process. In the production of aluminum, alumina (Al2 O3 ) is electrolytically reduced to aluminum in a cryolite (Na 3 ALF6 ) bath

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via the following reaction [48]: 2Al2 O3 + 3C → 4Al + 3CO2

(27.1)

The electrolytic refining process is performed in reduction cells, commonly known as smelting pots. An alumina reduction cell consists of a refractory lined steel shell upon which is placed a carbon lining, which serves as the cell’s cathode. Electrical current enters the cell through a bakedcarbon anode that is immersed in the cryolite. During the lifetime of an aluminum reduction cell, the carbon cathode becomes saturated with cryolite bath materials. Cyanide forms in the carbon cathode through a process that is not well understood but presumed to be caused by nitrogen from air diffusion on reacting with sodium and hot carbon [49]. The cyanides collect in the cooler areas of the potlining [14]. Each electrolytic cell has a predetermined operating life as the carbon lining is consumed and must be periodically replaced. The residual carbon lining, known as SPL, contains high concentrations of cyanide. The formation of cyanide in potlining is discussed further in Chapter 4. In the past, spent potliner material was typically used as onsite fill at many aluminum production facilities and resulted in cyanide impacts to soil and groundwater. Leachate from old SPL disposal sites typically has high pH (9–12) and elevated concentrations of a number of chemical species [13,14], most prominently fluoride (1200–8500 mg/l) and cyanide (150–4280 mg/l CNT ). The cyanide is primarily in the form of dissolved iron–cyanide species [50,51]. Table 27.2 provides ranges of cyanide species concentrations in aged leachate from legacy SPL disposal sites. The nature of these impacts is discussed further in Chapter 10. Today, SPL generated in the production of aluminum is a listed hazardous waste (K088) in the United States [52] and is managed as such. In a study of the subsurface fate and transport of cyanide in SPL leachate-contaminated groundwater at an aluminum production facility in North Carolina, Gilgore-Schnoor [51] examined a cyanide plume in fractured rock that derived from multiple SPL legacy-landfill areas onsite. The facility began operation in the early 1900s, and SPL storage onsite occurred from 1917 to the 1970s. SPL was used as fill material at a number of locations on the facility property. The site was underlain by 5 to 20 ft of residual soil, and below that was the fractured bedrock. As of the early 1990s, leachate from the SPL disposal areas had contaminated the groundwater in the unconfined aquifer comprising the residual soil and the fractured bedrock. From monitoring wells placed in the aquifer at the facility

TABLE 27.2 Cyanide Concentrations in Aged Leachate from Spent Potlining Disposal Sites Cyanide species Total cyanide Fe(CN)4− 6 WAD cyanide Free cyanide Amenable cyanide (CATC)

Concentrations, mg/l 150–4200 150–4280 0–20 20 0–150

Sources: Data from Blayden, L.C., Hohman, S.C., and Robuck, S.J., Proceedings of the Light Metals 1987, The Minerals, Metals and Material Society, Warrandale, PA, 1987; Kimmerle, F.W., Girard, P.W., Roussel, R., and Tellier, J.G., Proceedings of the Light Metals 1989, The Minerals, Metals and Material Society, Warrendale, PA, 1987.

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and hydraulically downgradient of the facility, it was determined that groundwater flow was away from the site toward a lake. Several distinct and overlapping areas of cyanide contamination were identified from review of monitoring data obtained from 1990 to 1993. Total cyanide concentrations in the groundwater ranged up to about 50 mg/l. The multiple source areas contributed to the formation of one main plume of cyanide contamination in the groundwater. From the data available it was concluded that over the three-year time period of observation, the total cyanide plume exhibited very limited lateral transport. Using plume center of mass analysis following the methodology of Freyberg et al. [53], an overall velocity of plume transport was calculated and compared to the average linear velocity of groundwater movement over the same time period. This analysis resulted in a cyanide plume retardation factor (groundwater velocity/cyanide plume velocity) of 1.5. This retardation factor greater than 1.0 indicates some retention of cyanide on aquifer material, perhaps by adsorption of iron–cyanide species, which are generally dominant in SPL leachate as discussed above. Considering the short observational period and other sources of uncertainty with the analysis, however, the strongest conclusion that can be reached is that the apparent velocity of cyanide transport at the site is similar to that of the groundwater. That is to say, the data are sufficient to indicate that the cyanide transport was not highly retarded. Cyanide had clearly moved into the groundwater beneath the site and then had been transported offsite, in about the same amount of time required for water to follow the same path. Sampling performed in the lake downgradient of the facility in the early 1990s periodically resulted in detectable concentrations of total cyanide in a part of the lake closest to the aluminum production facility. Concentrations of 5µg/l were measured on several occasions. In response, a systematic and carefully planned water quality study focused on cyanide species was conducted in the lake in 1996. No detectable concentrations of cyanide species were measured. Remediation activities have been undertaken at the site to try to reduce SPL leachate generation and mitigate cyanide transport in groundwater away from the site. Several of the SPL disposal areas were capped to reduce infiltration into the SPL and hence the volume of leachate generation. In addition, some sewers, which had been serving as inadvertent collectors of SPL-impacted groundwater through inflow and infiltration, have been slip-lined. Additional measures for addressing the subsurface cyanide contamination and the potential for discharge into the lake are under consideration.

27.3.4 ELECTROPLATING SITE REMEDIATION Cyanide solutions are used in many electroplating operations (Chapter 26). These solutions, along with solutions of acids, bases, and dissolved metals used in electroplating are kept in storage tanks onsite. Historic maintenance and secondary containment measures for these tanks were not what they are today, and as a result there is legacy subsurface contamination at many sites of former electroplating operations. These sites are often small in scale, but cyanide concentrations can be very high because of the high concentrations in the alkaline cyanide plating baths that were employed in these processes. Figure 27.3 shows a schematic view of a former electroplating site in Pennsylvania. As indicated on the diagram, a series of liquid storage tanks had been located outside the electroplating facility. The site overlies a shallow, unconfined aquifer perched on a layer of fine-grained material 12–15 ft below ground surface. A deeper unconfined aquifer existed 5–10 ft below the perched aquifer. A series of shallow wells were installed in the perched aquifer around the location of the former tanks and monitored in 1985–1986. The results of this monitoring, summarized in Table 27.3, revealed significant total cyanide concentrations, especially at locations closest to the former tanks. No analyses other than total cyanide were conducted on the samples collected. Some additional wells placed in the deeper aquifer revealed no detectable cyanide. The subsurface cyanide contamination detected in the perched groundwater at the former electroplating site had been in place for at least several years and probably much longer. It was concluded

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Site boundary — fence W1

W4

Alkaline waste

Acid waste

Acid rinse

Cyanide rinse

Cyanide waste

W3

W5 W2

Scale 10 ft

Electroplating facility

FIGURE 27.3 Schematic plan view of an electroplating storage tank site in Pennsylvania.

TABLE 27.3 Cyanide Concentrations Measured in Groundwater Monitoring Wells at a Former Electroplating Site in Pennsylvania (see Figure 27.1) Total cyanide (mg/l) Sample date May 1985 Nov 1985 Dec 1985 Jun 1986 Jul 1986 Aug 1986 Sep 1986

Well W1

Well W2

Well W3

Well W4

Well W5

NAa 45 NA NA 130 110 130

2300 1100 1400 2300 1530 1700 550

11 18 28 NA NA 11 3.9

NA NA 91 52 6.8 2.5 0.99

NA 180 290 805 845 830 905

a NA = not available.

Source: Data from D.A. Dzombak.

that active biodegradation of cyanide in the contaminated zones was unlikely to be significant because of the high concentrations of cyanide, the persistence of the cyanide, and the potential for much of the cyanide to be complexed with metals. Aqueous flushing via injection-recovery wells was considered as was in situ alkaline chlorination. Both of these remedial approaches were rejected because of the potential for pushing cyanide-contaminated water over the edges of the perched layer and down into the deeper, uncontaminated aquifer. The remedial approach selected was the emplacement of a passive collection trench filled with gravel in the area of the former storage tanks. The trench was designed to penetrate the perched aquifer (but not its lower confining layer) and be installed to a depth of 9 to 12 ft. A water-level-activated sump pump was provided in the trench to remove the water collected for subsequent treatment.

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27.4 SUMMARY AND CONCLUSIONS • Management of cyanide-impacted soil and groundwater requires a tiered approach involving: (i) establishment of a basis for compliance and the analytical methods for determining it; (ii) development of a detailed understanding of the concentration and distribution of the various cyanide forms at the site through careful choice of analytical methods; and (iii) development of a management decision framework in which the desired outcome is specified and the steps required to achieve that outcome are outlined. • A key factor in managing cyanide-impacted soil and groundwater is to match the appropriate cyanide analytical method with the cyanide species that forms the basis for the regulations of interest. In most cases, the cyanide species of primary interest is the toxic free cyanide (HCN, CN− ). • After agreement on the primary cyanide species of interest, compliance with the regulations can be evaluated and the need for remedial action can be determined. • Selection of the appropriate remedial action, which could but does not necessarily include application of active treatment technologies, will also depend upon the cyanide species that are the subject of concern at the site. • Significant experience has been gained with management of subsurface cyanide contamination at aluminum production facilities, electroplating facilities, hydrometallurgical gold mining facilities, and former manufactured gas plants.

REFERENCES 1. USEPA, Common chemicals found at Superfund sites, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, http://www.epa.gov/superfund/resources/chemicals.htm, accessed: March 22, 2005. 2. USEPA, Method 335.2: Cyanide, total (titrimetric, spectrophotometric), Rev. 1980, Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, http://www.nemi.gov, 1979. 3. USEPA, Method OIA-1677: Available cyanide by flow injection, ligand exchange and amperometry, EPA-821/R-99-013, U.S. Environmental Protection Agency, Office of Water, Washington, DC, 1999. 4. USEPA, Method 335.1: Cyanides amenable to chlorination (titrimetric, spectrophotometric), Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, 1979. 5. ASTM, Designation D 2036-95. Standard test method for weak acid dissociable cyanides in water, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 6. APHA/AWWA/WEF, Method 4500-CN: Cyanide, in Standard Methods for the Examination of Water and Wastewater, 20th ed., Clesceri, L.S., Greenberg, A.E., and Eaton, A.D., Eds., American Public Health Assoc., American Water Works Assoc., and Water Environment Federation, Washington, DC, 1998. 7. Zheng, A., Dzombak, D.A., Luthy, R.G., Sawyer, B., Lazouskas, W., Tata, P., Delaney, M.F., Zilitinkevitch, L., Sebroski, J.R., Swartling, R.S., Drop, S., and Flaherty, J., Evaluation and testing of analytical methods for cyanide species in municipal and industrial contaminated waters, Environ. Sci. Technol., 37, 107, 2003. 8. ASTM, Designation D 4282-95. Standard test method for determination of free cyanide in water and wastewater by microdiffusion, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 1998. 9. MDEQ, Table 1. Groundwater: residential and industrial-commercial Part 201 generic cleanup criteria and screening levels; Part 213 tier I risk-based screening levels, Michigan Department of Environmental Quality, Remediation and Redevelopment Division, http://www.deq.state.mi.us/documents/deq-rrdOpMemo_1-Attachment1Table1GW.pdf, accessed: April 22, 2005,

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10. Callahan, M.A., Slimak, M.W., Gabel, N.W., May, I.P., Fowler, C.F., Freed, J.R., Jennings, P., Durfee, R.L., Whitmore, F.C., Maestri, B., Mabey, W.R., Holt, B.R., and Gould, C., Water related environmental fate of 129 priority pollutants, Vol. 1, EPA-440/4-79-029a, U.S. Environmental Protection Agency, Office of Water and Waste Management, Washington, DC, 1979. 11. Kavanaugh, M.C., Deeb, R.A., Markowitz, D., Dzombak, D.A., Zheng, A., Theis, T.L., Young, T.C., and Luthy, R.G., Cyanide formation and fate in complex effluents and its relation to water quality criteria, Report 98-HHE-5, Water Environment Research Foundation, Alexandria, VA, 2003. 12. RETEC, Duck Spring pilot scale wetland monitoring project. Interim annual report for 2004, report prepared for Alcoa Remediation Operations, The RETEC Group, Pittsburgh, PA, 2005. 13. Blayden, L.C., Hohman, S.C., and Robuck, S.J. Spent potliner leaching and leachate treatment, in Proceedings of the Light Metals 1987, The Minerals, Metals and Materials Society, Warrendale, PA, 1987, p. 663. 14. Kimmerle, F.W., Girard, P.W., Roussel, R., and Tellier, J.G. Cyanide destruction in spent potlining, in Proceedings of the Light Metals 1989, The Minerals, Metals and Materials Society, Warrendale, PA, 1989. 15. Theis, T.L., Young, T.C., Huang, M., and Knutsen, K.C., Leachate characteristics and composition of cyanide-bearing wastes from manufactured gas plants, Environ. Sci. Technol., 28, 99, 1994. 16. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Nakles, D.V., Subsurface fate and transport of cyanide species at a manufactured-gas plant site, Water Environ. Res., 71, 1205, 1999. 17. Ghosh, R.S., Nakles, D.V., Murarka, I., and Neuhauser, E.F., Cyanide speciation in soil and groundwater at manufactured gas plant (MGP) sites, Environ. Eng. Sci., 21, 752, 2004. 18. Smith, A. and Mudder, T., The Chemistry and Treatment of Cyanidation Wastes, Mining Journal Books, Ltd., London, 1991. 19. Irwin, R.J., Cyanide(s) in general, in Environmental Contaminants Encyclopedia, National Park Service, Water Resources Division, Water Operations Branch, Fort Collins, CO, 1997. 20. Canter, L.W. and Knox, R.C., Ground Water Pollution Control, Lewis Publishers, Chelsea, MI, 1986. 21. Nyer, E.K., Groundwater Treatment Technology, 2nd ed., Von Nostrand Reinhold, New York, 1992. 22. LaGrega, M.D., Buckingham, P.L., and Evans, J.C., Hazardous Waste Management, 2nd ed., McGraw-Hill, New York, 2001. 23. Kastman, K.H. and Zimmerman, R.E., Cyanide waste disposal site neutralization, in Proceedings of Conference on Geotechnical Practice for Disposal of Solid Waste Materials, Am. Soc. Civil Engineers, New York, 1977, p. 831. 24. Ghosh, R.S., Dzombak, D.A., Luthy, R.G., and Smith, J.R., In situ treatment of cyanide-contaminated groundwater by iron cyanide precipitation, Water Environ. Res., 71, 1217, 1999. 25. Dzombak, D.A., Ghosh, R.S., and Luthy, R.G., Method for treating water contaminated with cyanide, U.S. Patent 5,837,145, Nov. 17, 1998. 26. Ebbs, S.D., Bushey, J.T., Poston, S., Kosma, D., Samiotakis, M., and Dzombak, D.A., Transport and metabolism of free cyanide and iron cyanide complexes by willow, Plant Cell Env., 26, 1467, 2003. 27. Bushey, J.T., Modeling cyanide uptake by willows for phytoremediation, Ph.D. thesis, Carnegie Mellon University, Pittsburgh, PA, 2003. 28. USEPA, Guidance for conducting remedial investigations and feasibility studies under CERCLA, EPA-540-G-89-004, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, Washington, DC, 1988. 29. USEPA, Abandoned mine site characterization and cleanup handbook, EPA-530-C-01-001 (also EPA-910-B-00-001), U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, Washington, DC, 2001. 30. Kosson, D.S., van der Sloot, H.A., Sanchez, F., and Garrabrants, A.C., An integrated framework for evaluating leaching in waste management and utilization of secondary materials, Environ. Eng. Sci., 19, 159, 2002. 31. MDEQ, Table 2. Soil: residential and commercial I Part 201 general cleanup criteria and screening levels; Part 213 tier I risk-based screening levels, Michigan Department of Environmental Quality, Remediation and Redevelopment Division, http://www.deq.state.mi.us/documents/ deq-rrd-OpMemo_1-Attachment1Table2SoilResidential.pdf, accessed: April 21, 2005,

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32. ASTM, Designation D 3987-85(2004). Standard test method for shake extraction of solid waste with water, in Annual Book of ASTM Standards, Vol. 11.02, ASTM International, West Conshohocken, PA, 2004. 33. MDEQ, Sampling and analysis. Attachment 2. Soil leaching methods, Michigan Department of Environmental Quality, Remediation and Redevelopment Division, http://www.deq.state.mi.us/documents/ deq-rrd-OpMemo_2_Attachment2.pdf, accessed: April 21, 2005. 34. MADEP, Background documentation for the development of an “available cyanide” benchmark concentration, Massachusetts Department of Environmental Protection, Boston, MA, http://www. mass.gov/dep/ors/files/cn_soil.htm, accessed: July 27, 2004, 1998. 35. USEPA, Method 1311: Toxicity characteristic leaching procedure, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 5, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. 36. USEPA, Method 1312: Synthetic precipitation leaching procedure, in SW-846: Test Methods for Evaluating Solid Waste: Physical/Chemical Methods, Rev 5, U.S. Environmental Protection Agency, Office of Solid Waste, Washington, DC, 1998. 37. USEPA, Method 335.3: Cyanide, total (colorimetric, automated UV), Methods for the Chemical Analysis of Water and Wastes, EPA-600/4-79-020, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Cincinnati, OH, http://www.nemi.gov, 1979. 38. MDEQ, Table 3. Soil: industrial and commercial II, III, and IV Part 201 generic cleanup criteria and screening levels; Part 213 tier I risk-based screening levels, Michigan Department of Environmental Quality, Remediation and Redevelopment Division, http://www.deq.state.mi.us/documents/deq-rrdOpMemo_1-Attachment1Table3SoilCommercial.pdf, accessed: April 21, 2005. 39. Fey, D., Desborough, G., and Finney, C., Analytical results for total digestions, EPA-1312 leach, and net acid production for twenty-three abandoned metal-mining related wastes in the Boulder River watershed, northern Jefferson County, Montana, OFR-00-114, U.S. Geological Survey, Denver, CO, 2000. 40. Yukselen, M.A. and Alpaslan, B., Leaching of metals from soil contaminated by mining activities, J. Haz. Mat., 87, 289, 2001. 41. NRC, Hardrock Mining on Federal Lands, National Academy Press, Washington, DC, 1999. 42. USEPA, Applicability of the toxicity characteristic leaching procedure to mineral processing waste, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, Washington, DC, http://www.epa.gov/epaoswer/other/mining/minedock/tclp.htm, accessed: April 19, 2005. 43. EPRI, Geochemistry, fate, and three-dimensional transport modeling of subsurface cyanide contamination at a manufactured gas plant, Final Report No. 1001301, Electric Power Research Institute, Palo Alto, CA, 2001. 44. Mudder, T., Miller, S., Russell, L., Cox, A., and McWharter, D., Introduction to the Biopass system, an alternative treatment process closure of spent heap leach pads, in Cyanide Monograph, T. Mudder, Ed., Mining Journal Books, Ltd, London, 1998. 45. Mudder, T., Miller, S., Russell, L., Cox, A., and McWharter, D., The Biopass system, Phase I: Laboratory evaluation, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 46. Guilfoyle, L. Treatment of cyanide in a heap leach pad, in Proceedings Mining into the Next Century: Environmental Opportunities and Challenges, Minerals Council of Australia, Townsville, 1999, p. 447. 47. Cellan, R., Cox, A., Uhle, R., Jenevein, D., Miller, S., and Mudder, T., The Biopass system, Phase II: full-scale design and construction, in Cyanide Monograph, Mudder, T., Ed., Mining Journal Books, Ltd, London, 1998. 48. Sanders, R.E., Aluminum and aluminum alloys, in Kirk-Othmer Encyclopedia of Chemical Technology, online edition [November 15, 2002], John Wiley & Sons, Inc., New York, 2002. 49. Haupin, W.E., Environmental considerations, Crit. Rev. Appl. Chem., 20, 176, 1987. 50. Dzombak, D.A., Dobbs, C.L., Culleiton, C.J., Smith, J.R., and Krause, D., Removal of cyanide from spent potlining leachate by iron cyanide precipitation, in Proceedings of WEFTEC96, Vol. 3, Part I. Remediation of Soil and Groundwater, Water Environment Federation, Alexandria, VA, 1996, p. 107. 51. Gilgore-Schnorr, G., Evaluation of subsurface fate/transport of cyanide and fluoride at spent potlining leachate field sites, M.S. thesis, Carnegie Mellon University, Pittsburgh, PA, 1994.

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52. USEPA, Land disposal restrictions; treatment standards for spent potliners from primary aluminum reduction (K088), Federal Register, 63(185), 51253, U.S. Environmental Protection Agency, Washington, DC, 1998. 53. Freyberg, D.L., A natural gradient experiment on solute transport in a sand aquifer. 2. Spatial moments and the advection and dispersion of nonreactive tracers, Water Resources Res., 22, 2031, 1986. 54. Kjeldsen, P., Behaviour of cyanides in soil and groundwater: a review, Water, Air, Soil Pollut., 115, 279, 1999.

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Index Acartia clausi, 268, 274, 276–277 Acetaldehyde, 107–108 Acetonitrile, 2, 10, 20, 37, 86, 124 O-Acetylserine, 105 Acidosis, 243 Acinetobacter, 98 see also Bacteria Acremonium strictum, 95, 103 see also Fungi Acrylonitrile, 41, 45, 49–51, 98, 451 Activated alumina, see Adsorption Activated carbon, see Adsorption Activated sludge, 459, 464–468 Acute toxicity see Toxicity, ecological see Toxicity, human Adenine, 2 Adsorption activated alumina, 414, 421 activated carbon, 413–418, 422, 430, 506, 524, 533 biosorption, 464 in groundwater transport, 196–198, 233 in municipal wastewater treatment, 504 ion exchange, 414, 418–422 on soil organic matter, 61, 202, 230–231 see also Free cyanide, adsorption see also Strong metal-cyanide complexes see also Weak metal-cyanide complexes After-effects, 241 Agriculture chemicals, cyanide in, 41, 43 see also Phytoremediation, design considerations Air stripping free cyanide removal, 413, 429–434, 464, 467, 504, 506, 533 see also Free cyanide, volatilization Air-water exchange, HCN Henry’s Law constant, 176–177 flux chambers, 178 mass transfer resistance, 177–178 vapor pressure, 176 volatilization rate, as half life, 176 see also Free cyanide, volatilization Alamosa River, 171 Alanine, 106–108 Alcaligenes faecalis, 102 spp., 95 xylosoxidans, 102 see also Bacteria Alfalfa, 26–27 Algae, 25, 32, 226 Alkaline chlorination, 63, 124, 393, 510, 575, 585 see also Chemical oxidation see also Thermal treatment technologies

Alkaline hydrolysis, 65 see also Thermal treatment technologies Almonds, 28 Aluminum smelting, 6, 16, 161, 191, 193, 429, 577–578 incidental production of HCN, 52–53, 226, 574 see also Groundwater see also Soil see also Spent potlining Ambient water quality criteria (AWQC), see Water quality criteria Americamysis, 266, 268, 274, 276–277 Amide, 98, 111 Amino acid, 4, 25, 27, 98, 102, 105–108, 110–111, 115, 495 α-Aminobutyric acid, 106–109 α-Aminobutyronitrile, 107–108 4-Amino-4-cyanobutyric acid, 108–109 γ -Cyano-α-aminobutyric acid, 106, 108 L-isoleucine, 27 L-leucine, 27 L-phenylalanine, 27 L-tyrosine, 27 L-valine, 27 Ammonia, 3–4, 6, 33, 37, 64, 96–101, 103, 107–109, 112, 115, 142, 429–430, 440, 450, 453, 460–462, 464, 466–467, 471–472, 522, 524, 530–531, 545–554, 562–563 Ammunition waste, 481 Amygdalin, 86–87, 124 Amyl nitrite, 245 Analytical methods, solids direct acid distillation, 156–157, 160, 162–163 see also Analytical methods, water see also Extraction methods Analytical methods, water and regulatory issues, 380–382 available cyanide, 125–126, 133, 324–325, 361, 381, 384, 512, 572–573, 577–578 cyanate, 142 cyanide amenable to chlorination, 72, 125–126, 132–134, 325, 361, 572–573, 577–578 cyanogen halides, 142–143 free cyanide by direct colorimetric development, 137–138 free cyanide by gas chromatography, 137 free cyanide by ion-selective electrode, 141 free cyanide by microdiffusion, 138–140, 161, 164, 278, 325, 512, 520, 573 interferences and pretreatment, 127–128, 141–142 metal cyanides by ion chromatography, 134–136, 161 metal cyanides by reversed-phase ion-pair partition chromatography, 136 method detection limits, 146–147, 510 organocyanides, 143–144 physiologically available cyanide, 345

591

CYWS

“L1666_C028” — 2005/10/12 — 18:18 — page 591 — #1

Index

592

Analytical methods, water (contd.) thiocyanate, 141–142 total cyanide by distillation, 125, 128–132, 156, 159, 164, 337, 361, 520, 572, 577–578 weak acid dissociable (WAD) cyanide, 125–126, 132, 325, 510, 512, 520, 573 Andrussow process, 42, 45 Animal arthropods, 36–37, 226 biosynthesis, 36 cyanide role in, 37 feed, cyanide in, 41 foals, toxicity incident, 230 tissue, analysis, 163–164 Antidotes, 245–246 Apoplasm, 488 see also Modeling, plant uptake see also Plant Applicable or relevant and appropriate requirements (ARARs), 372 Apoptosis, proteins BNIP3, 245 uncoupling protein, 245 Arabidopsis thaliana (Mouseear cress), 111, 482 see also Plant Arginine, 111–113 Arrowhead, common, 484 Arthrobacter spp., 95, 461 see also Bacteria Ascorbate, 112 Asparagine, 105, 107, 110–111 Aspartate, 105, 110–111 Assimilation, see Metabolism, pathways Aquarium fish, 219–222 Aquatic biota, 179 Aquatic plants dose, normalization, 288 exposure pathways, 287 freshwater, 267, 273 marine, 275 mechanisms of toxicity, 287 Aquatic-dependent wildlife, toxicity, 285–306 bioavailability, 286 exposure pathways, 286 route of exposure, birds, 299–305 route of exposure, mammals, 290–303 toxicity thresholds, 305–306 Atlantic salmon, see Salmo salar Atlantic silverside, see Menidia menidia Atmosphere, cyanide in, 5, 37–38, 209 Attached growth, 468–470 Auxin, 111 Available cyanide, see Analytical methods, water Avena sativa L. (Oat), 486–487, 491 see also Plant AVR process, 432, 533

Bacillus sp., 461 megaterium, 95, 104–105 pumilus, 95, 97 see also Bacteria

CYWS

Bacteria, 5, 34–35 assimilation, 94, 104–106 consortium, 99 cyanide tolerance, 97, 474 cyanogenesis, 25, 226 degradation, 94–101, 113, 229, 492 see also Free cyanide, biodegradation see also Metabolism, by see also Metabolism, pathways Barium cyanide, 81 Barley, see Hordeum vulgare L. Beans, 26–28 Beilby process, 44 Benzonitrile, 98, 102 Benthic organisms, 373–374 Berlin Green, 81, 84, 423, 426 Berlin White, 81, 84, 423 Best available treatment (BAT) technology, 11, 360, 363–364, 366, 368, 370, 372, 555 Best practical treatment (BPT) technology, 360, 363–364, 366, 368, 370, 372 Bioaccumulation, 305, 332, 335, 343, 346, 495 Bioavailability, 101, 115, 252–254, 286, 480, 490–491 complex cyanide, 253, 278, 304–305 dissolved oxygen, effect on, 254 free cyanide, 253, 278, 286 metallocyanides, 253, 286 pH, effect on, 254 simple cyanide, 304–305 temperature, effect on, 254 total cyanide, 278 see also Metabolism Biogas production, 98 Biomagnification, 332, 335 Bioassay, fish, 172–173 Biodegradation, see Free cyanide Bioleaching, 530–531 Biological treatment, 115, 388, 459–474, 506–507, 533, 552, 555, 560 contact time, 463 hydraulic retention time, 466–467 pH, 463 sludge age, 466–467 temperature, 463 Biomass burning, 6, 37, 209 Biomining, see Phytoextraction Biopass, 471 Biosorption, 467, 470 Birds, see Aquatic-dependent wildlife Bird’s foot trefoil, see Lotus corniculatus L. Blast furnace, 518, 522, 541–554 incidental production of HCN, 52 process, 541–544 sludge, 159–160 wastewater characteristics, 503 wastewater treatment, 545–554 Blood, human cyanide levels, 242 Bluegill sunfish, see Lepomis macrochirus Bonding cyanide ion, 57–58 metal-cyanide complexes, 58 weak metal-cyanide complexes, 65 Botanical, see Plant

“L1666_C028” — 2005/10/12 — 18:18 — page 592 — #2

Index

593

Brain lesions, 238 Fos protein, 239 sensitivity to cyanide, 316 Brassica Brassicaceae, 286 sp., 482 juncea (Indian mustard), 480, 483, 492, 496, see also Rhizofiltration napus (rape), 112 see also Plant Brevibacterium, 98 see also Bacteria Brook trout, see Salvelinus fontinalis Bullrush, river, 484 softstem, see Scirpus validus L. Burkholderia cepacia, 95 see also Bacteria Bur-reed, American, 484 Bush bean, see Phaseolus vulgaris L. β-glucosidase, 26

Cadmium cyanide, 19, 65–66, 70–72, 81, 98–99, 481 see also Weak metal-cyanide complexes Calcium cyanide, 21, 79, 81 Cancer irroratus, 268–269, 274, 276–277, 359 Candida guilliermondii, 95, 102 tropicalis, 103 see also Fungi Carbon dioxide product, 96–97, 100, 102, 112, 115, 461 metabolic cycle, 96, 108, 112 Carbonyl sulfide, 100 CAS, see Enzyme Case studies see also Electroplating see also Fate processes see also Hydrometallurgical gold mining see also Manufactured gas plant see also Spent potlining Cassava (Manihot esculenta), 26, 29, 226 see also Konzo disease Castner process, 44–46 Cattail, see Typha sp. Ceratophyllum demersum L. (coontail), 484 see also Plant Cesium cyanide, 81 Chemical oxidation, treatment technology alkaline chlorination, 393–398 electrolytic decomposition, 447–448 high temperature alkaline chlorination, 443–445 hydrogen peroxide, 398–404, 472, 533 Inco’s Air/SO2 , 406–408 ozone, 393, 398–404 photocatalytic oxidation, 404–406, 575 wet air oxidation, 450–452 Chemical plant and cyanide production, 6, 16 wastewater characteristics, 501, 503 wastewater treatment, 387

CYWS

Cherry, 26–27, 230 Chlamydomonas reinhardtii, 111 Chloramination, 508–509, 533, 540 Chloramphenicol, 104 Chlorinated solvents, 481 Chlorine, 63, 394–398, 443, 533 Chlorination, cyanide formation in, 507–512 Chromobacterium violaceum, 95, 104–106 see also Bacteria Chronic toxicity see Toxicity, ecological see Toxicity, human Citrobacter freundii, 95, 99, 104–105 see also Bacteria Citrulline, 113 Clean Water Act (CWA), 352, 382, 520, 537, 555, 559 see also Regulations, surface water see also Regulations, wastewater Coal coking, see Coke plant Coal gasification, 16, 52, 85, 494 Coastal waters, see Marine waters Cobalt cyanide, 19, 65, 73, 81, 129, 134–136, 174, 481 see also Strong metal-cyanide complexes Cobalt cyanate, 84 Cocklebur, 110 Cofactor, 97, 101–102, 109 Coke plant, 192, 462–464, 466–467, 471, 518, 522, 555–563 coke oven gas, 555 process, 555 incidental production of HCN, 52, 226 wastewater characterization, 85, 503, 517, 555–559 wastewater treatment, 103, 387–388, 429, 467, 559–560, 562 Community, ecological, see Phytoremediation, design considerations Comprehensive Environmental, Response, Compensation, and Liability Act (CERCLA), 351–352, 360, 362, 371–374, 382 Consortium, see Bacteria Constructed wetlands, see Phytoremediation, treatment Coontail, see Ceratophyllum demersum L. Copper cyanide, 17, 65–66, 70–73, 80–82, 98–99, 134–136, 172, 238, 253, 353, 355, 359, 415, 481 see also Weak metal-cyanide complexes Coral reefs, 219–222 see also Toxicity, ecological Corn, 26–27 Cosmetics, 16, 81, 482 Cottonwood, see Populus, deltoids L. Crab, see Cancer irroratus Cryptococcus humicolus MCN2, 95, 101–102 sp. UFMG-Y28, 102 see also Fungi Cyanamide, 112–113 Cyanate, 16–19, 61–63, 73, 79, 82–84, 96, 100–101, 103, 108, 111–112, 124, 142, 394, 398, 404, 406–407, 443, 447, 461, 508, 541 measurement, 142 Cyanidation process, 41, 524, 527–528, 530 Cyanide amenable to chlorination and regulations, 353, 359, 361, 380 see also Analytical methods, water

“L1666_C028” — 2005/10/12 — 18:18 — page 593 — #3

Index

594

Cyanide cycle natural cyanide cycle, 25, 225, 229–231 anthropogenic cyanide cycle, 231–233 Cyanide fishing, 210, 219–222, 335 Cyanide microcycle, 25, 96, 115, 492 Cyanide poisoning, see also Toxicity asphyxia, 241 effects, 239 symptoms, 240 Cyanisorb process, 432, 533 β-Cyanoalanine, 105, 107–111, 115 γ -Cyano-α-aminobutyric acid, 106, 108 Cyanobacteria, 32–33, 335 see also Algae Cyanocobalamin, 87 Cyanogen bromide, 16–17, 81–82 measurement, 143 Cyanogen chloride, 16–17, 63, 82, 124, 394, 508–509 measurement, 143 Cyanogenic bacteria, 25, 34, 104, 114, 199 see also Bacteria Cyanogenic fungi, 25, 34–35, 107, 114 see also Fungi Cyanogenic glycosides, 2, 16, 20, 25–27, 30, 36, 86, 108, 110, 114–115, 163–164, 234, 482, 491, 496 biosynthesis, 27 precursor amino acids, 27 Cyanogenic lipids, 26–27 Cyanogenic plants, 26–28 see also Plant Cyanohydrin, 86, 97, 124 Cystathionine, 106 Cysteine, 105, 110 Cytochrome c oxidase, 252, 287 Daphnia, 259, 264, 271 Defense mechanism, see Toxicity, plant Defoliant, see Herbicide Degradation see Free cyanide see Metabolism, pathways see Strong metal-cyanide complexes see Weak metal-cyanide complexes Denitrification, 466, 562 Detoxification, see Toxicity, ecological Dhurrin, 27, 86–87, 124 2,4-D (dichlorophenoxyacetic acid), 111 see also Herbicide Dioxins, 479 DNA gene sequencing, 96, 100, 102, 108, 111 nucleotide, 112 transgenic, 113, 481, 490 Doses, human dose-response, 316 sublethal, 239 lethal, 241, 318 Draeger tubes, see Free cyanide, volatilization Drinking water see MAC see MCL and MCLG see Regulations, drinking water see Regulations, international Dyes, 41, 81, 226, 482

CYWS

Earth’s atmosphere, 3, 5 Echinochloa crus-galli (barnyard grass), 111 see also Herbicide see also Plant Ecological risk assessment endpoints, 336–337 conceptual models, 115, 332–336 effects, 337–340 exposure, 343–345 problem formulation, 332–337 risk characterization, 346, 495–496 species mean acute values, 340–342 species sensitivity distribution, 340–342 stressor identification, 342 Effects, human on behavior, 239 on brain electrical activity, 239 on hearing, 239 on heart, 239–240 (see Heart) Elderberry, see Sambucus sp. Electron transport chain, mitochondria ATP synthesis, 109 cyanide-resistant respiration, 109–110, 482, 491, 495 NAD(P)H, 96–97, 104, 110 Electroplating, 6, 43, 191, 501, 518, 521, 535–541, 571 process, 535–537 process water treatment, 420, 427, 447–448 site remediation case study, 584–585 wastewater characteristics, 420, 501, 503 wastewater treatment, 387–388, 420, 427–428, 464, 471 Endodermis, 488 see also Modeling, plant uptake see also Plant Endpoints, assessment aquatic, 275–277, 336, 576 ecological, 337, 576 Enterobacter sp., 99 aerogenes, 104–105 strain 10–1, 105 see also Bacteria Environment Australia, 528, 530 Enzyme activity, 110–111, 482, 495 alkaline phosphatase, 495 alternative oxidase, see Electron transport chain amidase, 98, 101–103 γ -aminobutyric acid transaminase, 108 arginase, 113 asparaginase, 105 carbamoyl phosphate syntetase, 112 carbonic anhydrase, 112, 495 cyanamide hydratase, 109, 113 cyanase, 96–97, 109, 112 cyanidase (cyanide dihydratase), 98 cyanide hydratase (formamide hydratase), 97–98, 101–103 cyanide oxygenase, 96–97, 99, 101 β-cyanoalanine synthase (CAS), 104–108, 110–111, 482 β-cyanoalanine hydratase, 105, 111 γ -cyano-α-aminobutyric acid synthase, 104–106 cysteine synthase, 105

“L1666_C028” — 2005/10/12 — 18:18 — page 594 — #4

Index

595

cytochrome c oxidase, see Electron transport chain formamidase, 101 β-glucosidase, 26 glutamic acid decarboxylase, 108 hydroxylnitrile lyase, 26 linamarinase, 26 3-mercaptopyruvate sulfur transferase, 111 nitrate reductase, 110 nitrilase, 98, 108–109, 111 nitrile aminohydrolase, 111 nitrile hydratase, 98, 101 nitrogenase, 97 peroxidase, 33 processes, 94, 104, 108–109 rhodanese, 104, 107–108, 111, 242, 482 succinic semialdehyde dehydrogenase, 108 sulfurtransferase, 104, 106, 108, 111, 242, 482 thiocyanate hydrolase, 97, 100 thiosulfate:cyanide transferase, 111 Escherichia coli, 95, 97–100, 104–105 see also Bacteria Ethylene enzyme regulation, 110 synthesis, 16, 32–33, 37, 108, 110–111, 114, 491, 495–496 see also Auxin European Commission, 376, 377, 380 see also Water Framework Directive Exopolymers, 463, 467 Exposure, ecological, 335, 343–345 Exposure, human acute, 241 at former MGP sites, 310–311 chronic, 241–242 dietary, 316, 326 environmental, 179–188, 238 industrial, 238, 577 occupational, 238 sublethal, 239 Extraction methods, solids acid solution, 157–158 alkaline solution, 158–160, 164–165 effect of extraction time, 162–163 effect of liquid-solid ratio, 161–162 neutral solution, 158 quality control, 165 see also Animal, tissue see also Plant, tissue

Fate processes, surface waters adsorption, 173, 230–231 air-water exchange, 179, 184–185, 187, 252 biological, 172, 178–179, 185, 187, 231–232 case study, wetland, 182–188 factors, environmental fate determining, 171–176, 231–233 hydrolysis, 173 modeling, 179–188 photolysis, 179, 184, 187 precipitation, 173–174 see also Free cyanide, volatilization see also Photodissociation see also Transport processes

CYWS

Fathead minnow, see Pimephales promelas Fermentation, see Metabolism, pathways Ferric Ferrocyanide (FFC), 6, 10, 16, 22, 48, 192, 228, 233, 310, 315, 318–319, 333, 382 see also Prussian Blue Ferricyanide, 6, 19, 65, 73–79, 83, 102–103, 145, 173–176, 178–179, 184, 193–195, 211, 253–254, 311, 315, 414, 418, 425–426, 482 Ferrocyanide, 6, 10, 19, 41, 73–79, 83, 102–103, 134–136, 145, 161, 172–176, 178–179, 184, 193–195, 211, 253–254, 311, 315, 333, 414, 416–419, 421–422, 425–426, 440–441, 472, 482 assimilation, 113–114, 493 at former MGP sites, 203–204, 310–311 biodegradation, 99–100, 178, 482, 486, 493 phytoremediation, 480–483, 485–486, 488–491, 493 production, 48–49 solubility, 485 toxicity, 318–320 Fertilizer, 112, 495 Fires, forest, 26, 37–38, 238 Fire retardants, 16, 43 Five spot burned moth, see Zygaena trifolii L. Fluidized bed reactor, 469 Foods, 238 see also Cassava Forage, 30–31 Forensics analyses, 164 Formaldehyde, 2, 37, 408 Formamide, 96, 98, 101–103, 115 Formate, 64, 98, 103 Formic acid, 96, 101–102, 115 Foxtail, see Setaria sp. Free cyanide adsorption, 61–62, 157, 197, 413, 493, 505–506 assimilation, 96, 108–109, 481, 488–493, 495–496 at former MGP sites, 310, 312 bioavailability, ecological, see Bioavailability, ecological and photolysis, 174–176 and regulations, 354–359, 372–373, 375–380, 382, 383 biodegradation, aerobic, 63, 96–97, 101, 179, 199–200, 211–212, 218, 229–230, 233, 381, 460–461, 463, 486, 493, 501, 505–506 biodegradation, anaerobic, 96–97, 460, 462 HCN, CN− , 16–18 HCN formation, 3–5, 18, 58, 227 HCN dissociation, 17, 58, 227 hydrolysis, 64 in atmosphere, 209 in environment, 481 oxidation, 62–63 oxidation states, 63 ozonation, 63 phytodecomposition, 233, 480–481, 484, 488–490, 492–493 sedimentation, 484 volatilization, 60–61, 114, 138, 198–199, 211–212, 218, 228, 233, 381, 481, 484, 488–490, 493, 496–497, 501, 504, 506 see also Analytical methods, water see also Metabolism see also Phytoremediation

“L1666_C028” — 2005/10/12 — 18:18 — page 595 — #5

Index

596

Free cyanide (contd.) see also Toxicity, ecological see also Toxicity, human Freshwater, cyanide in, 171–188 see also Regulations, groundwater see also Water quality criteria Fungi, 5, 34–35 assimilation, 106–108 cassava spoilage, 107 cyanogenesis, 25 degradation, 94–95, 98, 101–103, 113, 492 fungicide, 112 pathogenic, 101, 107, 115 psychrophilic basidiomycete, 107–109 root rot, 107 snow mold, 107 white rot, 101 winter crown rot, 107 see also Metabolism, by see also Metabolism, pathways Fusarium sp. lateritium, 102 nivale, 107 oxysporum, 95, 102 solani, 95, 101–102 see also Fungi Gas chromatography, see Analytical methods, water Gene, see DNA Generation, in brain, 242 Gloeocercospora sorghi, 95, 101–102, 115 see also Fungi Glucose, 95, 100 Glucosinolate, 482 Glutamate, 104, 106, 108–109 Glutamic acid, 107, 109 Glutamine, 111–112 γ -Glutamyl dipeptide, see β-Cyanoalanine Glycine, 2, 104–106 Gold cyanide, 19, 34, 73, 81, 98–99, 134–136, 413–414, 416, 418, 530 see Strong metal-cyanide complexes Gold mining, 1, 6, 8, 41, 43, 171, 191, 226, 460, 463–464, 469, 471, 522–535, 571 see also Hydrometallurgical gold mining Grass, rabittfoot, 494 Groundwater effect of oxidation-reduction potential, 194–195 management of cyanide, 571–575 treatment, 480–481 see also Fate processes see also Regulations, groundwater see also Spent potlining see also Transport processes Guy Lussac, 2 Hazardous waste, 362, 371–372 see also Resource Conservation and Recovery Act see also Regulations, hazardous waste Health advisories, 353, 355 see also Toxicity, human Heart failure, 239

CYWS

irregularities, 239 bradycardia, 239 Helianthus annus L. (sunflower), 480 see also Plant see also Rhizofiltration Helminthosporium trucicum, 101 see also Fungi Henry’s Law constant, HCN, 60–61 see also Free cyanide, volatilization Herbicides, 6, 41, 43, 111–112, 226, 481 Hexacyanoferrate see Ferricyanide see Ferrocyanide see Potassium hexacyanoferrate Histidine, 33–34, 112 Homestake Mine, 469, 471, 524, 526 Hordeum vulgare L. (barley), 483, 486–487, 491 see also Plant Humic acid, 416, 509 Hydrogen cyanide, 1–6, 10 and ecological toxicity, 252, 286–287, 289, 305 and regulations, 351–352, 355, 363, 381, 385 gas, 2, 4–6, 16–17 production, 4, 42, 45–46 see also Analytical methods, water see also Fires see also Free cyanide see also Toxicity, ecological see also Toxicity, human Hydrogen peroxide, 400–403 Hydrometallurgical gold mining, 1, 6, 41 carbon-in-leachate process, 416, 531 carbon-in-pulp process, 416, 430, 531, 533 cyanide recovery, 430–434, 523–524, 531–533 heap leach pad closure case study, 581–582 process, 523–424 resin-in-pulp process, 419 wastewater and treatment, 387–388, 417–419, 523–535 Hydroponic, see Phytoremediation Hydroxycobalamin, 88, 246 Hypochlorite, 63, 394

Impoundment ponds, 522, 524, 527, 533 see also Tailings ponds Incinerator, cyanide formation in, 53, 507, 509–511 Indian mustard, see Brassica juncea Indole, acetamide, 111 acetonitrile, 111 Ingestion, 29, 32, 238 Inhalation, 238, 312–313, 572 Inhibition, biological treatment, 460–462, 464–465 see also Metabolism Inks, 16, 41, 81, 482 In-plant controls, 518, 521–522 Intoxication, see Toxicity, ecological in vitro, see Metabolism in vivo, see Metabolism Ion chromatography, see Analytical methods, water Ion exchange, 531 see also Adsorption, separation technology Ion selective electrode, see Analytical methods, water

“L1666_C028” — 2005/10/12 — 18:18 — page 596 — #6

Index

597

Iron and steel production, see Steel production Iron cyanide adsorption, 197 mobility, 201–204 pH-pE diagram, 195, 197 see also Berlin Green see also Berlin White see also Ferric Ferrocyanide see also Ferricyanide see also Ferrocyanide see also Metal-Metal solids see also Precipitation, separation technology see also Prussian Blue see also Prussian Brown see also Strong metal-cyanide complexes see also Turnbull’s Blue α-ketobutyrate, 107 Ketoglutyric acid, 109 Kidney, 241 Klebsiella oxytoca, 95, 97, 99 see also Bacteria Konzo disease, 29–30, 241–242, 335 see also Cassava see also Toxicity, human Leptosphaeria maculans, 95, 101 see also Fungi Lacustrine system, 180 Land disposal restrictions, 371, 373 Landfill, 371–373 Lima beans, 26–28 Linamarin, 26, 86 Leach tests see also Extraction methods see also SPLP test see also TCLP test Lepomis macrochirus, 253, 255, 262, 270, 272, 276 Lesions brain, 238–239 Liver, 241 Lixiviants, 521, 530–533 see also Hydrometallurgical gold mining Lotus corniculatus L. (bird’s foot trefoil), 28, 230, 495 Lowest observed adverse effect level (LOAEL), 381 Lycopersicum esculentum L. (tomato), 483 see also Plant Lysine, 112 MacArthur-Forest process, 41–42, 524 see also Cyanidation process Mammals see Aquatic-dependent wildlife see Toxicity, ecological Manufactured gas plant (MGP), 1, 6, 10, 16, 191–192, 227, 310, 333, 571, 574 closure case study, 579–580 cyanide species concentrations, 310–313 leachate, 74, 161, 193 risks, human, 323–324 soils and wastes, 99, 113, 157–163, 192, 387, 482, 495, 577 soil treatment, 453, 472

CYWS

Marasmius, 107 see also Fungi Marine waters, cyanide in coastal waters, 211–218, 510 see also Cyanide fishing see also Water quality criteria Maximum acceptable concentration (MAC), free cyanide, 376–377 Maximum contaminant level (MCL), free cyanide, 354–355, 358–359, 372, 376, 380 Maximum contaminant level goal (MCLG), free cyanide, 355, 372, 380 Metabolism by rhodanese, 242 by sulfurtransferase, 242 Mechanisms, toxicity acidosis, 243 altered calcium balance, 243 neurotransmitter release, 243 oxidative stress, 243–244 Menidia menidia, 268, 274, 276–277 Mercury, 481, 496 Mercury cyanide, 65–66, 81–82, 133 see also Weak metal-cyanide complexes Metabolism carbon source, 95–96, 100–101, 104, 115 conditions, 95, 99–103, 107–108, 115, 491 co-contaminants, 103, 115 energetics, 104 inhibition, 99, 104, 109, 112, 482, 495 in vitro, 96, 104, 484 in vivo, 104, 113, 495 nitrogen source, 95–96, 99–104, 112, 115 sulfur source, 100 toxicity, 103 see also Plant, assimilation Metabolism, by, see Enzyme Metabolism, pathways assimilation, 95, 104–105, 108–115, 491 cyanogenesis, 96, 104, 108, 114–115 decarboxylation, 107 degradation, 95–104, 108, 115 fermentation, 100 hydrolytic, 96–98, 100–101, 105, 107–108, 110, 114–115 oxidative, 96–98, 100–101, 103, 115 nucleophilic, 112 reductive, 96–98, 115 substitution/transfer, 96, 105–106 synthesis, 106–108, 112 see also Enzyme Metal-cyanate complexes, 18, 84 Metal-cyanide complexes see Weak metal-cyanide complexes see Strong metal-cyanide complexes Metal-cyanide solids simple cyanide solids, 16, 21, 79–80 metal-metal cyanide solids, 16, 21–22, 80–84, 155, 157, 160, 192, 423–424, 462 solubility, 80–84, 192, 228, 311, 315, 462 Metal finishing, see Electroplating Metal-thiocyanate complexes, 18, 85 Methane, 3, 6, 98, 556 Methionine, 104

“L1666_C028” — 2005/10/12 — 18:18 — page 597 — #7

Index

598

Methyl Cyanoacetate, 98 Microdiffusion, free cyanide measurement by, see Analytical methods, water Microorganisms see Algae see Bacteria see Fungi Mill’s Crow process, 432 Miller, Stanley, 3, 555 Modeling assimilate, 489–490 conceptual, for ecological risk analysis, 333–334 cyanide species adsorption, 61–62 engineering application, 487, 493–494 fate and transport, 179–188, 312–313 MINEQL+, 67, 75 MINTEQ, 66–67, 74 nutrient solution, 113, 485 plant uptake, 114, 480, 486–490, 497 Molasses, 95, 99, 468 Mouseear cress, see Arabidopsis thaliana see also Plant Municipal wastewater characteristics, 501–502, 504, 506 cyanide sources, 502 cyanide formation, 502, 507–512 sludge incineration, 53 treatment, 504–507 see also Regulations, wastewater Murder, 237 Muskgrass, 496 Mysidopsis, see Americamysis

NAD(P)H, see Electron transport chain Necrosis, proteins BNIP3, 245 uncoupling protein, 245 Nickel cyanide, 10, 17, 19, 65–69, 81, 98–99, 102–103, 134–136, 155, 172, 194, 253, 481 see also Weak metal-cyanide complexes Nicotinamide, 111 Nitrate, 103, 472, 563 Nitrification, 461–464, 466, 467, 471, 506, 562–563 Nitriles assimilation, 105, 111–113 biodegradation, 98, 101–102 Nitrite, 103 Nitrogen metabolic cycle, 96, 108, 110, 112, 482, 485, 491, 497 Nitrosation, 512 No observed adverse effect level (NOAEL), 317–318, 381 Norcardia rhodochrous, 95, 98 see also Bacteria Nucleotide, see DNA

Oat, see Avena sativa L. Oceans, cyanide in, 209 Oil refinery see also Petroleum refining see also Wastewater discharges Omai, Guyana, 8

CYWS

Oncorhynchus mykiss, 257–258, 262–263, 270, 272, 276–277 Origin of life, and cyanide, 2–5 Organic carbon, soil, 413–414, 506 Organocyanides, 16, 18, 20, 86–88, 124–125, 335, 509 measurement, 143–144 Ornithine, 112 Outer space, cyanide in, 5 Oxidation see Chemical oxidation see Free cyanide see Strong metal-cyanide complexes see Weak metal-cyanide complexes Oxidation state cyanide, 63 cyanate, 63, 82 thiocyanate, 64, 86 Oxide box waste, 310 see also Manufactured gas plant Oxygen, 246, 406–408, 461–462, 466, 468–469 Ozone, 63, 398, 403 see also Chemical oxidation

Palladium cyanide, 73 see also Strong metal-cyanide complexes Parkinsonism, 241 Pea, see Pisum sativum Penicillum miczynski, 95, 101–102 see also Fungi Perca flavescens, 257, 270, 276–277 Permeable reactive barrier, 575 Pesticide, 16, 41, 226, 479 Petroleum refining, 16, 43, 518, 563–567 and selenocyanate, 231 process, 563–564 wastewater characteristics, 211–216, 218–219, 501, 503 wastewater treatment, 565 Pfluger, E., 2, 25 Pharmaceuticals, 16, 41, 43, 81, 226, 482 Phaseolotoxin, 112 Phaseolus vulgaris L. (bush bean), 482 see also Plant Phenol, 95, 103, 466–467, 522, 545–554, 560 Phenylalanine, 33 Philiota sp. adipose, 106 aurivella, 107 praecox, 107 see also Fungi Photodissociation, 76, 98, 111, 113, 174–178, 198, 211–212, 228, 233, 381, 404–406, 486, 493, 504 Photosynthesis, 495 Phytodegradation, 480–481, 492 see also Phytoremediation Phytoextraction, 199, 480, 492 see also Phytoremediation Phytoremediation benefits, 480, 485 design considerations, 480, 483–484, 487, 490–494 biomass management, 480–481, 490, 493 hydroponic, 114, 483, 485–486, 488, 493–494

“L1666_C028” — 2005/10/12 — 18:18 — page 598 — #8

Index

599

limitations, 480, 485, 490–495 metabolism, 111, 113–114, 480–482, 484–486, 492, 496 regulatory concerns, 480–481, 484, 490, 495–497 transport, 111, 113–114, 480, 484–485, 487–488, 490–492, 496 treatment, 112, 115, 480, 484, 486, 492–495, 575 see also Metabolism see also Plant Phytostabilization, 480 see also Phytoremediation Phytovolatilization, 480–481, 496 see also Phytoremediation Picrate paper, 114, 164, 497 Pigments, 16, 41, 44, 226 Pimephales promelas, 252, 256–257, 262, 270, 272 Pisum sativum (pea), 114, 496 see also Plant Plant age, 486 analysis, 113–114, 163–165 assimilation, 108–114, 484–486 biomass, 481, 483, 485, 490 cyanide-resistant respiration, see Electron transport chain cyanide role and impact, 27–32, 229 cyanogenesis, 25, 27–28 phloem, 490 physiology, 487–488 senescence, 111 stress, 110, 112, 490–492, 495–496 survey, 520–521 tissue, analysis, 163–165 tolerance, 111, 113, 482–483, 491 transformation processes, 108 transgenic, see DNA transpiration, 114, 482–483, 487, 490, 496 see also Aquatic plants see also Auxin see also Ethylene synthesis see also Phytoremediation Platinum cyanide, 19, 73, 129 see also Strong metal-cyanide complexes Pollution prevention, 518–519, 527 Pollution Prevention Act, 360–361, 519 Polychlorinated biphenyl (PCB), 479 Polycyclic aromatic hydrocarbon (PAH), 479 Polysulfide, 19, 64, 85, 448–450, 530–531 Pondweed, see Potamageton spp. Poplar, Balsam, see Populus trichocarp Populus sp. deltoids L. (cottonwood), 496 trichocarp (balsam poplar), 111, 483, 492 see also Plant Potamageton spp. (pondweed), 484 see also Plant Potassium cyanide, 2, 21, 79, 81–82, 113, 144, 155, 158, 164, 211–212, 238, 241, 353, 355, 496 Potassium hexacyanoferrate, 79, 99, 113, 160, 164, 212 see also Ferricyanide see also Ferrocyanide see also Hexacyanoferrate Potato, 110, 112

CYWS

Precipitation in groundwater transport, 196 separation technology, 228, 413, 423–428, 524, 531, 566 Pretreatment standards, see Regulations, wastewater Proprionaldehyde, 107 Proprionitrile, 98 Protein-mediated uptake, see Modeling, plant uptake Prussian Blue, 2, 10, 16, 41, 48, 79–82, 84–85, 103, 155, 165, 192, 228, 310, 382, 383, 385, 389, 423–428, 482–483 dissolution, 193–195 precipitation, 196 Prussian Brown, 81, 84, 423 Pseudomonas, 98–99, 104–105, 461 sp. aeruginosa, 95, 100 fluorescens, 95–98, 100 putida, 95, 99 nonliqurfaciens, 95 stutzeri, 95, 100 syringae, 112 see also Bacteria Publicly Owned Treatment Works (POTW) see Municipal wastewater see Regulations, wastewater Puget Sound, 511 Quinclorac (quinolinecarboxylic acid), 111, 482 see also Herbicide Rainbow trout, see Oncorhynchus mykiss Reactive cyanide, 373, 383 Recycle and reuse, 518, 522, 541, 545 Regulations, drinking water, 1, 149, 325–326, 355, 358–359, 372, 376, 380 Regulations, groundwater, 359–360, 362, 371–372, 377–378 Regulations, hazardous waste, 362, 371–373, 429, 537–539 Regulations, international, 375–380 Australia, 377–379 Austria, 378–379 Canada, 376–377, 379, 385 Denmark, 376–379 European Union, 376–377, 380 France, 378–379 drinking water guidelines, 354, 358, 359, 376, 383 Germany, 376–379, 386 Netherlands, 11, 311, 378–379 United Kingdom, 376–377, 379, 386 World Health Organization, 376–377 Regulations, sediment, 372–376, 378, 380, 383–385 Regulations, soil, 10–11, 114, 324–325, 372–373, 378–379, 572–573, 577 Regulations, surface water, see Water quality criteria Regulations, wastewater Clean Water Act, 149, 520, 537, 555, 559 coke plant, 505 discharge permits, 360, 371, 511 Pollution Prevention Act, 519 pretreatment standards for existing sources, 360, 363–364, 366, 368, 370, 538

“L1666_C028” — 2005/10/12 — 18:18 — page 599 — #9

Index

600

Regulations, wastewater (contd.) pretreatment standards for new sources, 360, 363–364, 366, 368, 370 pretreatment standards for POTWs, 503–505, 510–511 specific industries, 518, 524, 537–539, 545, 555, 564 steel production, 505 Resource Conservation and Recovery Act (RCRA), 351–352, 360, 362, 371–374, 382, 538, 578 Rhizofiltration, 480 see also Phytoremediation, treatment Rhizopus oryzae, 107 see also Fungi Rhizosphere, 480 Rhodococcus sp., 95, 98 see also Fungi Rhodopseudomonas gelatinosa, 97 see also Bacteria Rhizoctonia solani, 107 see also Fungi Rhizopus oryzae, 95 see also Fungi Rice, 110–111 Ricinine, 111 Risk and management of soils, 572, 575 ecological, 10, 275–277, 345–346 human, 10, 310, 320–324 Road salt, cyanide in, 16, 44, 193, 482, 503 Rotating biological contactor (RBC), 468–469 Rubidium cyanide, 81

Safe Drinking Water Act (SDWA), 352, 355 see also Regulations, drinking water Salix (willow), alba (white), 111 eriocephala (diamond), 111, 113–114, 482–483, 485–486, 491, 495–496 viminalis (basket), 111, 482 see also Plant Salmonella typhimurium, 105 see also Bacteria Salmo salar, 258, 270, 276 Salvelinus fontinalis, 258–259, 270, 272, 276 Sambucus sp. (elderberry), 30, 111, 495 see also Plant San Francisco Bay, 210–218 Scheele, Carl, 2, 237 Scirpus validus L. (Bullrush, softstem), 484, 495 see also Plant Scytalidium thermophilum, 95, 101–102 see also Fungi Sedimentation, cyanide removal in, 504–506 Sediment quality criteria, 374–375, 378, 383–384 see also Regulations, sediment Selenium, 480–481, 494 Selenocyanate, 231, 470, 494, 496, 566 Separation technologies, 388 see also Adsorption see also Precipitation Sequencing batch reactor, 464–465 Sequestration, 480 Serine, 105

CYWS

Setaria sp. (foxtail), 483 see also Plant Sewage sludge, see Metabolism Silver cyanide, 19, 65–66, 81–82, 98–99, 134–136, 253, 468, 481 see also Weak metal-cyanide complexes Silver cyanate, 82, 84 Sodium cyanide, 1, 6, 21, 41–48, 79, 81–82, 145, 155, 158, 160, 220, 335, 482, 533 Sodium hexacyanoferrate, 6, 79 see also Ferricyanide see also Ferrocyanide see also Hexacyanoferrate Sodium nitrite, 245 Sodium thiosulfate, 245 Soil management of cyanide, 571, 575–579 see also Iron cyanide see also Manufactured gas plant see also Metabolism see also Regulations, soil see also Spent potlining Sorghum, 26–29, 226, 230 sp., 115 bicolor L. (wild cane), 483, 486–487, 491 see also Plant Soybean, 110–111 Spent potlining, 161, 429, 577–578 leachate characterization, 227, 482, 583 leachate management, 577–578, 583–584 leachate treatment, 100, 387–388, 425–427, 441–445, 484 sludges and solids treatment, 445–446 see also Aluminum smelting Spills, 8–9, 171, 527, 533 SPLP (Synthetic Precipitation Leaching Procedure) test, 158, 577–578 Stable isotope, 111, 113, 484–487, 491, 496 Steel production and cyanide production, 16, 191, 226 wastewater characteristics, 501 see also Blast furnace see also Coke plant Stemphylium loti, 95, 101, 107 see also Fungi Strong metal-cyanide complexes adsorption, 76–78, 157, 416, 421–422, 493, 502, 505–506 and biodegradation, 461 bonding, 73 definition, 18–19 degradation, 95, 98–102, 113–114, 486, 493 dissociation, 75–76, 99, 103, 113, 481, 488, 490, 493 equilibrium constants, 74 formation, 18–19, 73, 227–228, 481, 572 oxidation, 78–79 phytoremediation, 479–480, 484–486, 492–493 see also Ferricyanide see also Ferrocyanide see also Photodissociation see also Phytoremediation Succinate, 105 Succinic semialdehyde, 107–109 Succinonitrile, 98

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Index

601

Sugarcane, see Molasses Sublethal effects, ecological, 271–273, 339–340 blood hormones, 271–273 feeding behavior, 272 gill damage, 271 gonadosomatic index, 272 locomotor behavior, 272 Suicide, 237, 241, 318 Sulfide, 17, 64, 103, 127, 141, 530, 565 Sulfurtransferase, see Metabolism, by Summitville, Colorado, 8, 171 Sunflower, see Helianthus annus Sweet flag, 484

Tailings ponds, 527, 529 see also Hydrometallurgical gold mining see also Impoundment ponds TCLP (Toxicity Characteristic Leaching Procedure) test, 157, 577–578 Temperature effect on HCN dissociation, 59–60 effect on HCN volatilization, 61 effect on HCN hydrolysis, 64 Terrorist attack, 237 Tetracyanonickelate (TCN), see Nickel cyanide Thalia, 494 Thallium cyanide, 81 Thermal treatment technologies, 388 calcium polysulfide, 448–450 electrolytic decomposition, 447–448 high temperature alkaline hydrolysis, 440–443 incineration, 445–446 thermal desorption, 452–453 thermal oxidation, 453 thermal plasma, 453 Thiobacillus sp., 530 thiocyanoxidans, 95, 100 thioparus, 95, 100 see also Bacteria Thiocyanate, 15–19, 30, 61–64, 84–86, 124, 127, 141–142, 316, 337–338, 395, 397, 407, 449, 461, 464, 466–467, 472, 508, 522, 524, 533, 555, 560, 562, 563 degradation, 95, 97, 99–101, 103, 466, 492 glucosides, 482 measurement, 141–142 metabolic product, 104, 106, 111, 115 toxicity, 252, 271, 286–287, 338–340, 482 Thiosulfate, 19, 64, 85, 111, 200, 530 Thiourea, 530 Tisza River, 8–9, 171 Tobacco, 113 Tomato, see Lycopersicum esculentum L. Total cyanide and regulations, 352, 354–359, 361–369, 371–375, 378–380, 382–383 see also Analytical methods, water Toxicity, ecological acute, 10, 252, 255–261, 268–269, 271–272, 274, 338–339, 343 avian, 332 behavior, 338

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chemical defense, 96, 101, 104, 108, 114–115, 482, 495 chronic, 266, 272–275, 288, 337 coral reefs, 219–222, 332, 335 detoxification, 94, 105, 107–108, 111, 114–115, 462, 491 fish, freshwater, 255–259, 266, 271–273 fish, marine, 266, 268, 273–275, 332 intoxication, 104, 108, 114, 482, 485, 491, 495 invertebrates, freshwater, 259–261, 266, 273 invertebrates, marine, 266, 268–269, 274–275 metabolic basis, 337 poisoning, 112, 230 reproduction, 339–340 soil invertebrates, 337 thyroid, 340 vitellogenin, 340 see also Aquatic-dependent wildlife, toxicity see also Bioavailability see also Sublethal effects, ecological Toxicity, human acute, 10, 26, 241, 316 chronic, 241–242, 316 death, 241, 243 Konzo disease, 29–30, 241–242, 335 see also Effects Transformation see Metabolism, by see Metabolism, pathways Transport processes groundwater, 200–204, 233, 572 surface waters, 173, 179–182, 184, 233 Treatment technology selection, 11, 387–389, 522–523, 575 Trichloroethylene (TCE), 481 Trichoderma polysporum, 95, 101–102 see also Fungi Trickling filter, 199, 459, 468, 472 Triticum aestivum L. (wheat), 482 see also Plant Turnbull’s Blue, 79, 81–82, 84, 155, 192, 194, 423–426, 482–483 Turtle, snapping, 495 Typha sp. (Cattail), 484, 494 see also Plant Ubiquinone, see Electron transport chain, cyanide-resistant respiration Uranium, 480 Urea, 108, 112–113 UV disinfection, cyanide formation in, 512 UV light, and metal-cyanide dissociation, see Photodissociation Vitamin B12 , see Cyanocobalamin and Hydroxycobalamin Volatilization see Air stripping see Air-water exchange see Free cyanide Waste ammonia liquor, see Coke plant Waste load allocation, 188

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Index

602

Wastewater characteristics, 389, 502, 504, 506, 518, 520 management, 510–512, 517–518, 526–535, 540–541, 545–554, 556–563, 564–566 see also Best available treatment (BAT) technology see also Best practical treatment (BPT) technology Wastewater discharges, cyanide in industrial, 210–216, 218–219 mining, 99, 111 municipal, 210, 216–219 petroleum refinery, 211–216, 218–219 Water Framework Directive, European Commission, 376–377, 380, 384 Water quality criteria, cyanide and effluent limits, 210, 502, 510 freshwater, acute and chronic, 1, 11, 251–252, 254, 285, 352–355, 359–360, 372, 573, 577 marine, acute and chronic, 11, 210, 251–252, 273, 275–278, 285, 352–355, 359–360, 510–511 see also Water Framework Directive Water effects ratio, 511 Weak-acid-dissociable cyanide see Analytical methods, water see Weak metal-cyanide complexes Weak metal-cyanide complexes adsorption, 68–71, 157, 505–506 bonding, 65 definition, 18–19

degradation, 95, 98–99, 101–102, 230 dissociation, 99, 103, 481, 502 ecological toxicity, 345 equilibrium constants, 66 formation, 18–19, 65, 221–228, 481, 575 oxidation, 71–73 toxicity, 278, 286 treatment, 389, 460–461, 524, 533, 552, 562, 563, 575 Wheat, see Triticum aestivum L. Wild cane, see Sorghum bicolor L. Wildlife, 533 Williamson’s Salt, see Berlin White Willow, see Salix Wine, cyanide in, 16, 44 World Health Organization, 376–377, 385 Yeast, 102 Yellow perch, see Perca flavescens Yellow Prussiate of Soda, 233 see also Sodium hexacyanoferrate Zero discharge, 521–522 Zinc cyanide, 17–19, 65–67, 70–72, 80, 81, 98–99, 172, 253, 419–420 see also Weak metal-cyanide complexes Zygaena trifolii L. (five spot burned moth), 28, 230, 495

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