biodegradable polymers from industrial biomass waste

Bio-Nanotechnology for Sustainable Environmental Remediation and Energy Generation

Bio-Nanotechnology

for Sustainable Environmental Remediation and Energy Generation

E. González, E. Forero (Eds.)

Bio-Nano Convergence Network 2016

Academia Colombiana de Ciencias Exactas Físicas y Naturales & Nanoscale Science and Technology Center Copyright © 2016 All rights reserved. No part of this book may be reproduced in any form without written permission from publishers. Design, diagramation and art by UniversoNano Media Design ISBN: 978-958-9205-90-7 Printed in Bogotá, Colombia. DISONEX. Editors of this Book Edgar E González Geophysical Institute, Faculty of Engineering, Pontificia Universidad Javeriana, Bogotá, Colombia Nanoscale Science and Technology Center Academia Colombiana de Ciencias Exactas Físicas y Naturales Enrique Forero Academia Colombiana de Ciencias Exactas Físicas y Naturales Colegio Máximo de las Academias de Colombia Bogotá, Colombia Cbionano-Fealac Convergence Network www.fealac.org www.cbionano.org Contact: María Isabel Loaiza Internacionalization Office COLCIENCIAS, Bogotá, Colombia

TABLE OF CONTENTS Bio-Nanotechnology: Challenges and Opportunities.......................... 17 Edgar González, Ivan Montenegro

Heavy Metals Contamination Heavy Metal Distribution in Mine Water at Firefly Village, Shikoku, Japan ................................................................................................... 33 Katsuro Anazawa

Arsenic in drinking water: Current situation and technological alternatives for removal....................................................... 39 Ma. Teresa Alarcón, Alejandra Martín, Liliana Reynoso, M. Piña-Soberanis

Modelling of Mercury Transport, Fate and Transformation in Continental Surface Water Bodies........................................................... 65 Nelson Obregón, Leonardo García, Diana M. Muñoz

Remediation Nano modified clays, bioclays and bio-leaching for water and sediments remediation .......................................................................... 103 N. Porzionato, L. M. Guz, M. Olivelli, G. A. Curutchet, R. J. Candal

Biotechnological Synthesis of Silver Nanoparticles using Phytopathogenic Fungi and their Microbicidal Effect ..........................135 Raquel Villamizar

Ecotoxicology Ecotoxicology in Nanotechnologies ......................................................

153

Andrea Luna-Acosta

Sustainable Energy Biorefinery by the hand of the nanotechnology: biodegradable polymers from industrial biomass wate................... 207 José Vega-Baudrit , Michael Hernandez-Miranda, Rodolfo González-Paz, Yendry Corrales-Ureña

Environmental and bioenergetic context of livestock production in the Comarca Lagunera region, Mexico............................................... 215 Luis A. Hernández, José L. González, Juan Estrada

CO2 from waste to resource: Conceptual evaluation of technological alternatives for its exploitation.................................... 233 Jorge Chavarro

Community action Community criteria for integral management of domestic solid waste ........................................................................................................ 249 Alejandro Martinez, Luz E. Muñoz, Erika Nadachowski

AUTHORS Alejandra Martín-Domínguez

Instituto Mexicano de Tecnología del Agua (IMTA)

Alejandro Martinez

Centro de Educación para el Desarrollo Corporación Universitaria Minuto de Dios

A. González-Herrera.


Katsuro Anazawa

Department of Natural Environmental Studies Graduate School of Frontier Sciences The University of Tokyo

Andrea Luna-Acosta

Department of Ecology and Territory Faculty of Environmental and Rural Studies Pontificia Universidad Javeriana

Diana M. Muñoz

Geophysical Institute, Faculty of Engineering Pontificia Universidad Javeriana

Edgar E González

Geophysical Institute, Faculty of Engineering, Pontificia Universidad Javeriana Nanoscale Science and Technology Center Academia Colombiana de Ciencias Exactas Físicas y Naturales

Erika Nadachowski

Sistema general de Areas protegidas, Coorporación Autónoma General del Risaralda CARDER 9

Gustavo Andrés Curutchet

Instituto de Investigación e Ingeniería Ambiental, CONICET Universidad Nacional de San Martín

Iván Montenegro STI Policy Unit, Colciencias Jorge Chavarro Centro de Investigaciones CENIGAA José Vega-Baudrit

Laboratorio Nacional de Nanotecnología CONARE-CeNAT-LANOTEC Laboratorio de Polímeros, POLIUNA, Universidad Nacional

José L. González Barrios

Instituto Nacional de Investigaciones Forestales, Agrícolas y Pecuarias. Centro Nacional de Investigación Disciplinaria en Relación Agua-Suelo-Planta-Atmósfera.

Juan Estrada Avalos

Instituto Nacional de Investigaciones Forestales, Agrícolas y Pecuarias. Centro Nacional de Investigación Disciplinaria en Relación Agua-Suelo-Planta-Atmósfera.

Leonardo García

Basic Science Department Universidad Jorge Tadeo Lozano,

Liliana Reynoso Cuevas

Centro de Investigación en Materiales Avanzados, S.C. (CIMAV) Unidad Durango, México

Lucas Martín Guz


Luis A. Hernández

Instituto Nacional de Investigaciones Forestales, Agrícolas y Pecuarias. Centro

Nacional de Investigación Disciplinaria en Relación Agua-Suelo-Planta-Atmósfera.

Luz E. Muñoz

Corporación Universitaria de Santa Rosa de Cabal

Ma. Teresa Alarcón-Herrera

Centro de Investigación en Materiales Avanzados, S.C. (CIMAV) Unidad Durango, México

M. Piña-Soberanis


Melisa Olivelli


Michael Hernandez-Miranda


Natalia Porzionato


Nelson Obregón

Geophysical Institute, Faculty of Engineering Pontificia Universidad Javeriana

Raquel Villamizar

Departamento de Microbiología, Facultad de Ciencias Básicas Universidad de Pamplona

Roberto Jorge Candal


Rodolfo González-Paz


Yendry Regina Corrales-Ureña


PREFACE The Forum for East Asia- Latin American Co-operation (FEALAC) is a conference where twenty (20) countries from Latin America and sixteen (16) from East Asia currently participate. It places its main focus the strengthening of political, cultural, educational, social, economic, scientific and technological relations amongst member countries. Both Colombia and Japan have recently been appointed cochairs of the Work Group on Science, Technology, Information and Education. Preliminary work led by the joint action of the Colombian Presidential Agency of International Cooperation (APC), the Administrative Department of Science, Technology and Innovation -COLCIENCIAS-, the Ministry of Foreign Affairs, and the Colombian Association for the Advancement of Science, resulted in the development of Phase 1 within the consolidation of the Network of scientific-technological convergence. As a result of this preliminary work, an outline of the main task was drawn for Phase 2, hence, allowing it to set the Network of Bio-Nano Convergence (Cbionano-FEALAC), which was developed under the institutional coordination and execution of COLCIENCIAS, with the support of the Ministry of Foreign Affairs, APC and ACAC. As a result of all the activities carried out in phases I and II, the main aspects that guide the development of this initiative were agreed. From international workshops held, the working groups proposed as components to be addressed within the framework of cooperation projects between countries that are part of FEALAC: i) Detection, measurement, monitoring of heavy metals into water; ii) Nanoremediation; iii) Bioremediation; iv) Biorefineries. Two key areas of convergence were identified once problems concerning the member countries were reviewed: environment and

energy. These areas play an important role in research, innovation and development policies due to their social impact and relevance to member countries.Some initiatives, encouraging open and wide consultation with relevant experts and key members of the community in order to identify the main issues that modern societies within the 21st century have to cope with at a global level, show that there is consensus when considering that issues regarding energy, water and the environment are part of the key challenges that must be addressed urgently and effectively in order to guarantee the planet’s sustainability. The proposal consists in addressing environmental issues placing the focus on contamination produced by heavy metals in fresh water resources. All countries within FEALAC are to a greater or lesser extent affected by this complex problem, which in addition to causing serious damage to the environment, also compromises food safety and the health of the population exposed to this sort of contaminant. Hence, in the specific case of contamination produced by arsenic on fresh water for human use in South and East Asia, some 50 million people are exposed to concentrations of arsenic with values higher than those recommended by relevant environmental and health authorities. In Latin America, it is estimated that some 5 million people are exposed to contamination produced by this metal. Governments of a large number of countries affected by contamination by heavy metals have formulated programs and policies oriented to obtain information and prepare mitigation and remediation plans regarding the presence of heavy metals. In Colombia, due to the significant problem that contamination by mercury poses at present, the Ministry of Environment and Sustainable Development launched a Unified National Plan providing clear guidelines on transfer of technology, promoting the use of clean technologies, encouraging training, and building awareness on the use of mercury and products containing it, in a drive to minimize its

impact and protect public health and the environment from its effects. On the other hand, the Colombian Nanoscience and Nanotechnology Network has established as one of its priority tasks the need to tackle contamination by heavy metals from the perspective of nanoscale technologies, specifically by means of measurement, monitoring, mitigation and nanoremediation. Despite the fact that within FEALAC countries there has been a number of initiatives aiming at monitoring, mitigating and remediating heavy metals present in fresh water for human consumption, there is a call to increase cooperation and joint research in an effort to address this serious environmental problem. Some countries lack information regarding the degree of contamination by this sort of contaminant and the size of the population exposed to it in excess of recommended values remains unknown. Bio-Nanotechnology offer new avenues for detection, measurement, monitoring, and remediation. There is no doubt that progress in detection, measurement and monitoring has been achieved by means of this technology. Within many FEALAC countries there is capacity to develop low-cost, high precision portable processes and systems. From a remediation point of view, one of the most important contributions that have sprung from the revolution that nanotechnology entails has been the production of nanomaterials, which involves, in turn, innovative and exceptional properties relevant to the completion of this sort of tasks. Likewise, bio-refinery offers an important opportunity to live up to energy challenges with a high degree of sustainability and environmental commitment, in tandem with the valuable contributions made by biomaterials when applied - amongst other uses- to heavy metal remediation. It is convenient that remediation tasks as well as energy production are framed in a single holistic context, in which strategies for remediation may be set up in conjunction with energy production, maintaining an optimal balance between results and the impact caused on the environment and living beings in general. This

methodology demands that the toxicological effects and life cycle of processes and nanomaterials used must continuously be assessed. It is beyond question that by means of the execution of this sort of projects, FEALAC leadership as a forum for the actual economic and political integration of member countries will be strengthened, hence contributing in the development of scientific knowledge and the application of technology, benefiting, as a result, both regions. The FEALAC-Cbionano International workshop Environment and Energy: Challenges and Opportunities from Bio and Nanotechnology was co-hosted in Bogotá Colombia by the Ministry of Foreign Affairs, Presidential Agency of International Cooperation (APCColombia), the Administrative Department of Science, Technology and Innovation –COLCIENCIAS, Colombian Association for the Advancement of Science -ACAC, Pontificia Universidad Javeriana and Universidad de los Andes. The international workshop was attended by 350 researchers in energy and environment, academic authorities, public and private functionaries, entrepreneurs, teachers and advanced students in the area of bio-nanotechnology, among others. The topics covered in the workshop were: • Detection and measurement of heavy metals in water. • Bio and nanoremediation of heavy metals in water. • Biorefinery, a strategic route to address the energy problem. This book contains some of the presentations made at the workshop, as well as contributions by experts from FEALAC countries on energy and environment from the opportunities offered by bionanotechnology.

“For human societies to achieve a productive, healthful, and sustainable relationship with the natural world, the public and private sectors must make environmental considerations an integral part of decision making” The National Academies of Sciences, Engineering, and Medicine

E. González, E. Forero (Eds) Bio-Nanotechnology for Sustainable Environmental Remediation and Energy Generation. ACCEFYN&NanoCiTec, Bogotá, 2016.

Bio-Nanotechnology: Challenges and opportunities Edgar González1, Iván Montenegro2

T

aking advantage of the capabilities offered by biology and the enormous potential of nanotechnology, in a convergence context, a promising scenario which adopts the name of bionanotechnology is being constructed. In this scenario, solutions of great impact and sustainability to address the environmental and energy problems can be proposed and developed. In this chapter, we take into consideration the main challenges facing society in the 21st Century and the ways in which biotechnology offers to address them in a context of sustainability, are presented. In addition, the governance of international cooperation in I & D in the areas of energy and environment, are analysed.

Geophysical Institute, Faculty of Engineering, Pontificia Universidad Javeriana. Bogotá D.C., Colombia. Nanoscale Science and Technology Center, Bogotá D.C., Colombia. e-mail: [email protected] 2 STI Policy Unit, Colciencias, Bogotá D.C., Colombia, e-mail: [email protected] 1

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Bio-Nanotechnology

Introduction Nanoscience and nanotechnology are oriented to the study and manipulation of matter and energy at the nanoscale, where fundamental processes and components that support the structure and behaviour of all existing nature take place. When atoms and/or molecules are associated to form entities with close to nanometer dimensions, the physical and chemical behaviour of these entities –called nano-objects- are very sensitive to their composition, shape and size [1]. Properties such as electrical conductivity, elasticity, heat capacity, dispersion and light absorption, within many others, are drastically modified by changes in the aforementioned aspects. This makes the nanoscale behavior of matter a novel approach of great importance for potential applications and uses. Taking advantage of the capabilities offered by biology and the enormous potential of nanotechnology, in a context of convergence, a promising scenario which adopted the name of bio-nanotechnology is being constructed: “atom-level engineering and manufacturing using biological precedents for guidance” [2]. In this scenario are being developed strategies and scientific and technological tools to address the major challenges facing society in the 21st century. Between the main challenges the environmental problem and energy sustainability are highlighted. The crisis in water quality as well as the urgent need to develop methodologies and systems with the capacity to produce clean energy that fulfil the criteria of efficiency and sustainability, are some of the problems that can be tackled by bio-nanotechnology. However, to achieve this goal, it is necessary to increase international cooperation and strengthen mobility programs to transfer all this knowledge to new generations. 20

Challenges and opportunities

The role of bio-nanotechnology in the 2st century The substantial developments in science and technology that took place during the twentieth century, marked the consolidation of a society based on knowledge, but that is drastically affected by many problems that compromise, among others, environmental, energy, water and agricultural sustainability. As a result, a number of initiatives have been undertaken to face these challenges: •

At the request of the National Science Foundation, the National Academy of Engineering has published a list of the 14 grand challenges for engineering in the 21st century “considered essential for humanity to flourish” [3]. These challenges have been selected on the basis of opinions gathered from experts and from the general public around the world since 2006. These challenges include, among others, the need to make solar energy economical, provide access to clean water, develop carbon sequestration methods, manage the nitrogen cycle, and pass over engineer better medicines.

•

The Millennium Project was founded in 1996 by the Smithsonian Institution, Futures Group International and the United Nations University as an independent nonprofit global participatory research think-tank of futurists, scholars, business planners, and policy makers who work for international organizations, governments, corporations, NGOs, and universities [4]. Fifteen global challenges have been identified which provide a framework to assess the global and local prospects for humanity. Some of these challenges are: sustainable development and climate change, clean water, energy, science and technology.

•

To replace the Millennium Development Goals, the United Nations launched in 2015 the 2030 Agenda for Sustainable 21

Bio-Nanotechnology

Development and the 17 Sustainable Development goals. These goals are integrated and indivisible balancing the three dimensions of sustainable development: economic, social and environmental [5]. In the goals are included essential aspects such as the availability and sustainable management of water and sanitation, sustainable and modern energy for all, sustainably managed forests, to fight against climate change and its impacts, to protect, restore and promote sustainable use of terrestrial ecosystems, to combat desertification, to cease and reverse land degradation, and to halt biodiversity loss. •

The Royal Geographical Society has established a programme to analyse the biggest social, environmental and economic challenges facing the UK in the coming decades. The initiative falls within the context of the 21st Century Challenges. The list of challenges includes: low carbon energy, climate change, sustainability, air pollution, food security and water security, among others [6].

In order to respond to the challenges identified in the different initiatives listed above, emerging technologies such as bio and nanotechnology can play strategic roles. Figure 1 shows the most significant challenges that can be confronted with the use of these technologies. For example, the problem of clean water -which is understood in all initiatives as one of the main challenges-, is being looked at from the point of view of the strengths and opportunities offered by new processes, nanomaterials and devices that will allow the improvement of monitoring, mitigation and remediation activities [7]. The steady population growth, demand for food, industrial development, inadequate waste-water purification and treatment as well as the rise in pollution sources, are some aspects that are 22


Figure 2. Some of the challenges that can be assumed from the bionanotechnology. Reproduced with permission of NanoCiTec.

creating a true “water crisis”. According to the World Health Organization, 780 million people lack access to safe drinking water. This number is increasing exponentially, opening the way for a severe crisis of survival. The UN warns that, with the current level of consumption, 60 per cent of the world’s population will suffer water shortages by 2050. Adequate water management and implementation of effective sanitation programs to achieve the goals of the 2030 Agenda for Sustainable Development are urgent 23

Bio-Nanotechnology

and necessary. With the use of new materials it will be possible to develop portable sensors with high sensitivity and low cost, which will enable the configuration of networks for monitoring and measuring contamination of water bodies [8-9]. This, in turn, will help to determine causes of pollution, mobility of pollutants and the impact on the environment and on living creatures. Furthermore, it will be possible to develop bio-nanostructured membranes and filters to purify contaminated waters efficiently and safely. By using microorganisms it is possible to create nanostructured systems for wastewater treatment and energy production from degradation of organic matter found in the wastewater. An holistic approach to produce clean energy from environmental sanitation is one of the biggest challenges that must be faced. (Figure 2).

Figure 2. Holistic approach: to produce clean energy from environmental sanitation. 24


Information on scale about production of nanomaterials, life cycle, costs and impact on the environment and on living things is still lacking.

Responsible bionanotechnology: an essential challenge The supply of new materials is growing exponentially, as their use for consumer products. By 2014, a number close to 18000 products with nanotechnology were identified [10]. However, studies on the impact and risks of using nanomaterials are still incipient as it is illustrated in figure 3, where the number of articles published by year is shown.

Figure 3. Trendlines in the number of publications on nanomaterials, risk and exposure. Modified image from [10].

Life cycle assessments, risk factors for living things, and ecotoxicity of bio-nanotechnological products, are some of the main challenges to be considered in the context of emerging technologies to solve environmental and energy issues. Finding answers to these 25

Bio-Nanotechnology

questions will ensure the appropriate use of bio-nanotechnologies in order to make sure that the solutions provided now do not become problems to be solved in the future.

Converging technologies The so-called converging technologies -biotechnology, nanotechnology, information technology and cognitive technologies- have attracted considerable attention because of their importance in consolidating the initiatives that constitute the roadmap of innovation and development on this era. Transcending the use of the concept of interdisciplinarity as a result from a convergence nano-bio-info-cogno (NBIC) [11], these technologies are in the forefront of the efforts to respond to the challenges faced by the agendas of environmental and energy sustainability, food security, education and health, among others. The concept of new technologies in the context of convergence has given rise to a series of debates and studies by different sectors of society seeking transformative solutions that will have a positive impact on the quality of life. One of the main definitions of unification of NBIC originated in the need to understand the social implications of nanoscience and nanotechnology, derived from several meetings and workshops supported by the National Science Foundation and the National Science and Technology Council of the United States. The meeting, which took place in New York in 2004, turned out in the publication of a book under the name of Managing Nano-BioInfo-Cogno innovations: Converging Technologies in Society. [12]. In this book, it is stated that emerging technologies will play an important role in the road to progress and development of society as well as in the identification of new business models. The convergence finds its inspiration in the “unity of nature at the nanoscale”, where all sciences find common ground, as well as in 26


the significant achievements in the implementation and capacity of manipulation and control of matter at the nanometer scale. The European Commission, under the Sixth Framework Programme and in the context of the project Converging Technologies and Their Impact on the Social Sciences and Humanities, leads the issue of the role of converging technologies beyond the posture of the new science unit and nanoscale reductionism. According to the report published in 2008, in the heart of the new concept of converging technologies “are relations, synergies or fusions between broad fields of research and development, such as nanoscience and -technology, biotechnology and the life sciences, information and communication technologies, cognitive science and neurotechnologies. Robotics, Artificial Intelligence and other fields of research and development (R&D) are also taken into account in the discussions” [13].

Gobernance of international cooperation in I&D Environment energy and their applications based on nanotechnology and biotechnology are related to and conceived as global challenges due to the fact that the spreading of these problems has increased the sources and geographic extent of such challenges. Governance is conceived as the interaction of state and nonstate actors in the area of STI cooperation, coordination and policy making. These actors engage in governance as the process of defining principles, rules, regulations and decision-making procedures [14]. With regard to an environmental context and the relationship between science and policy it is useful to take into account epistemic and legal institutions. The former involving functions 27

Bio-Nanotechnology

such as research and development, technology transfer aimed at providing solutions to environment problems. Legal institutions are the ones empowered to make legal rules and adopt cooperative policies. Furthermore, the way in which states organize the process of creating shared scientific knowledge pertinent to environmental cooperation is referred to as epistemic cooperation, and the international institutions involved in creating shared scientific knowledge are considered epistemic institutions [15]. Regarding the case of independent epistemic institutions in a context in which states have an incentive to naturally coordinate environmental policies because the costs and benefits are primarily felt domestically, the technology transfer process becomes a key issue in the sense that actors have a clear incentive to adopt commercially viable and environmentally-friendly technologies. When collective action is necessary to address a global environmental problem, hierarchy is the optimal relationship between legal and epistemic institutions; hierarchy is the best means of ensuring the availability of a scientific record that is credible to the states bargaining in international legal institutions. The hierarchical degree of the relationship between epistemic and legal institutions depends on asset specificity –the extent to which information has value to multiple regulators–, credibility of the scientific record, and costs of governance [15]. On the other hand, one of the key dimensions of governance is sharing knowledge and intellectual property management. Undoubtedly, the Intellectual Property System, IPS, has encouraged innovation by helping to overcome market failures. In concrete terms, the Intellectual Property Rights, IPR, are not an impediment to the use and dissemination of technologies in so far as apart from being national in scope and limited in 28


time, patent holders can decide to license, to whom to do it and under what terms. Nonetheless, the IPR system shows at least some limitations as far as it does not generate enough incentives for innovation because it can create “patent thickets”, blocking patents, and “patent trolls” [16]. A patent thicket can be defined as an overlapping ensemble of patent rights which delay innovation because it requires innovators to reach licensing contracts for multiple patents from multiple sources. Patenttroll firms purchase patents and then they enforce said patents against purported infringers without themselves intending to manufacture the patented product or supply the patented service. In the field of nanotechnology, patent thickets have emerged limiting the sharing and the use of critical knowledge, impeding, in turn, downstream innovation, and preventing the development of more complex technologies due to exorbitant transaction costs. Hence, patenting nanotechnologies actually reduces commercial competition by making the use of some nanotechnologies highly expensive. Furthermore, a number of nanotechnology patents cover basic science in the quantum field, which raises serious doubts about the ownership of science [17].Some lessons learned and some conclusions on governance and international cooperation can be highlighted. In regards to sharing knowledge and IP management, independent epistemic institutions have a wide and specific demand for scientific and technology records and for specialized information related to the use of technology. Actors have a clear incentive to adopt commercially viable and environmentally-friendly technologies, whose content is channelled by means of these institutions. Dependent epistemic institutions contribute to overcome problems of uncertainty and collective actions, because they give developing countries the opportunity to oversee the R&D process and to resolve the uncertainty problems; and the legal institutions resolve the action’s collective problem establishing legal rules binding to all the affected parties. 29

Bio-Nanotechnology

Strengthening R&D in nanotechnology, biotechnology, their convergence and their applications could be supported under different governance frameworks of international STI cooperation. Insofar as these knowledge and technology fields are related to energy and environmental remediation actions whose goals and impacts are clearly perceived in every country, ongoing projects or new ones should be better linked to independent epistemic institutions such as universities, and public or private research centers. Alternatives and complementations to the traditional IPR system are justified, aiming to encourage innovations such as a prize fund approach, a congressional bill introduced in United States Congress, and the WHO Report about health needs in developing countries [16]. A free, open-source technology capacity building and development approach, and an innovative IPS, supported by a sound governance framework, can create a virtuous cycle, since they provide real opportunities of accessing knowledge, which would in turn lead to strengthening scientific and high-end technology skills.

Acknowledgement Thanks to Vicerrectoría de Investigaciones, Pontificia Universidad Javeriana (PPTA: 5126).

References [1] González E. The new era of nanomaterials. J. Nano Sc. Tech. 2013, 1, 84. [2] Goodsell D. Bionanotechnology Lessons from Nature. Wiley-Liss. Canada, 2004. [3] http://www.engineeringchallenges.org/

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[4] http://www.millennium-project.org/millennium/challenges.html [5]https://sustainabledevelopment.un.org/post2015/transformingourworld [6] https://21stcenturychallenges.org/ [7] Gonzalez, E.; Marrugo, J.; Martinez, V. (Eds.) El problema de Contaminación por mercurio. Oportunidades y capacidades desde la nanotecnología. RedNanoColombia, Bogotá, 2015. [8] Salinas S.; Mosquera N.; Yate L.; Coy E.; Yamhure G.; González E. Surface Plasmon Resonance Nanosensor for the Detection of Arsenic in Water. Sensors & Transd. 2014, 183, 97. [9] Reyes C.; Coy E.; Yate L.; González E.; Nanostructured and selective filter to improve detection of arsenic on surface plasmon nanosensores. ACS Sensors 2016, DOI: 10.1021/acssensors.6b00211 [10] González E. Nanomateriales, Riesgos y Sostenibilidad en: La Influencia de internet, genética y nanotecnología en la medicina y el seguro. U. Externado, 2015, ISBN 9789587724035. [11] Echeverra J. Interdisciplinariedad y convergencia tecnocientífica nano-bioinfo-cogno. Sociologías. 2009, 11 (22), 22-53. [12] Sims, W.; Roco, M. Managing Nano-Bio-Info-Cogno innovations: Converging Technologies in Society. Springer, Netherlands, 2006. [13] Andler, D.; Barthelmé, S.; et al. Converging Technologies and their impact on the Social Sciences and Humanities (CONTECS) An analysis of critical issues and a suggestion for a future research agenda Final Report. 2008. [14] OECD. Meeting Global Challenges through Better Governance: International Co-operation in Science, Technology and Innovation. OECD Publishing, 2012. [15] Meyer T. Epistemic Institutions and Epistemic Cooperation in International Environmental Governance. Transnational Environmental Law, 2013,15, 44. [16] Stiglitz J.; Greenwald B. Creating a Learning Society: A New Approach to Growth, Development and Social Progress. New York: Kenneth J. Arrow Lecture Series, 2014. [17] Pearce J. Open-Source Nanotechnology: Solutions to a Modern Intellectual Property Tragedy. Nano Today, 2013, 339-341.

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Heavy metal distribution in mine water

Heavy metals Contamination

33


Heavy metal distribution in mine water at Firefly Village, Shikoku, Japan Katsuro ANAZAWA

R

iver water and river sediments were collected from downstream area of an abandoned copper mine. This region is known for its clean water environment, and heavy metal concentration in downstream water of the mine was as low as the background level of the region. The river sediments, however, contained high concentration of heavy metals. This phenomenon was understood that when mine water with low pH is neutralized by river water with high pH, dissolved heavy metals are precipitated and concentrated in sediments. The thermodynamic simulation showed that a neutralization treatment could possibly perform 80100 % removal of heavy metals from the aqueous phase.

Department of Natural Environmental Studies, Graduate School of Frontier Sciences, The University of Tokyo. 5-1-5 Kashiwanoha, Kashiwa Chiba 277-8563, Japan. e-mail: [email protected]

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Introducción Japan was once a world’s leading mining nation, and now has nearly 5,000 abandoned mines. Among them, mine pollution such as water pollution has occurred at about 450 sites. [1] In many cases, mine owners are missing, and in those cases, responsibilities for pollution are unclear. Occasionally, those abandoned mines are hidden in the mountains, and water pollution sneaks up on the downstream unawares. In this study, water and sediments of a contaminated river by mine water were investigated to understand geochemical behaviors of heavy metals, and to propose a countermeasure to reduce the pollution at a typical mountain village.

Study area and analytical methods Misato district of Yoshinogawa city in Shikoku Island is recognized as a “firefly village” with clean water environment (Figure 1). In this district, Genji-Botaru (Luciola cruciata), which is designated as a national protected firefly species, are observed in countless numbers from the end of May to early in June. Strangely enough, no firefly is observed along a main river flow in the district. Various reasons for the lopsided geographical distribution of fireflies have been discussed, such as vegetation or toxicity of agricultural chemicals. Among those potential factors, heavy metals emitted from an abandoned copper mine existing at the upper stream area of the river were considered as the most suspicious cause. However, low concentration of heavy metals was found in the river water, which was considered harmless to the environment. [2] In order to clarify the cause of the firefly locality, geochemical investigation was performed in this district. Water and sediment were collected from the rivers in the district for chemical analysis. Sample collection was performed 4 times between 2002 and 2007,

36


Figure 1. Location map of the sampling sites.

and the samples were taken from 16 fixed points with several additional points each time. Water temperature, pH and electrical conductivity (EC) were determined at the sampling points. Dissolved heavy metals in river water were determined by atomic absorption spectrophotometry (AAS). Heavy metals in sediments were determined by AAS after mixed acid digestion of HF, HClO4 and HNO3. Determination of mercury was performed by cold-vapor AAS.

Results and discussion The analytical results are almost identical to the previous work. [3] The mine water and the river water contained less than 0.1 mg/ dm3 for Cd, Pb, As, Ni, Cr, Co and Mo, and less than 1.0 ng/dm3 for Hg. On the other hand, extraordinary high concentration of Cu and Zn was found in the mine water (sampling point A in Figs. 1, 2) with a magnitude 3-σ above the average concentration in this study area (Figure 2). The heavy metal concentrations in the 37

Bio-Nanotechnology

Figure 2. Variation of dissolved copper (Cu) and zinc (Zn) in river water.

downstream river water are almost equivalent to the background level. For sediments, among 10 heavy metals, only Cu and Zn also showed higher concentration in mining area than the regional background level, with a magnitude of one sigma (>1σ) above the average in the study region. In contrast to the river water case, the concentrations of Cu and Zn in the sediments show the highest values at point H1, not at point A (Figure 3). Point H1 is the first sampling point downstream of the junction of the mine water flow and the river. This phenomenon was quantitatively understood by stoichiometry considering the volume fraction of the mine water flow and the main river as 17: 83. [3] Dissolved Cu and Zn in the mine water with low pH were precipitated by the mixing with high pH river water and were included into the river sediments. Environmental threat to the 38


Figure 3. Variation of copper (Cu) and zinc (Zn) in sediments

fireflies’ survival was provoked not by water pollution, but by high concentration of Cu and Zn in the river sediments. As a proposal of countermeasure technique to reduce the outflow of heavy metals in the mine water, thermodynamic simulation was performed on the basis of neutralization treatment with calcium carbonates (CaCO3). The possible precipitates by neutralization reaction were assumed to be Cu4(OH)6SO4 for copper, Fe2O3 for iron, MnO2 for manganese and ZnSiO3 for zinc. Under consideration of reaction efficiency at the confluence of the mine water flow and the main river (point H1), 10% amount of solid CaCO3 was assumed to be eluted into river water as Ca2+. Saturation indices of minerals including heavy metals were calculated on the basis of WATEQ4F database. [5] The calculation results showed that a simple neutralization tank with CaCO3 will possibly be capable to perform 77-100 % removal of the heavy metals (Table 1). 39

Bio-Nanotechnology

Table 1. Removal efficiency and the composition change of the heavy metals in the mine water by neutralization treatment (theoretical value1)) [4]

Solution Metal Untreated water (mg/dm3)

Treated water (mg/dm3)

Ca

52.97

91.02

Cu

2.64

0.59

Fe

0.19

Mn Zn 1) 2)

Precipitation Removal efficiency (%)

Chemical Formula

(ton/yr)

CaCO3

-81.02)

77

Cu4(OH)6SO4

3.10

0.00

100

Fe2O3

0.23

1.16

0.00

100

MnO2

1.57

3.77

0.04

99

ZnSiO3

6.88

--

[Ca ][CO3 ] = 0.1 Ksp (calcite) Dissolution 2+

2-

References [1] JOGMEC. Mine Pollution Control. JOGMEC NEWS 2013, 35, 16. [2] Shinomura Y.; Anazawa K.; Sato M. The relationship between the river environment and the habitat of firefly in Misato, Tokushima Prefecture, West Japan. Environmental information science Extra, Papers on environmental information science. 2005, 297-302. [3] Anazawa K.; Kaida Y.; Shinomura Y.; Tomiyasu T.; Sakamoto H. Heavy-metal distribution in river waters and sediments around a “firefly village”, Shikoku, Japan: Application of multivariate analysis. Anal. Sci. 2004, 20, 79-84. [4] Anazawa K.; Shinomura Y.;Tomiyasu T. The behavior of heavy metals in mine water and the proposal of countermeasure technique at “firefly village”, MIsato, Yoshinogawa City, Tokushima Prefecture, Japan. Environmental information science Extra, Papers on environmental information science. 2007, 21, 601-606. [5] Ball J.; Nordstrom D. User’s manual for WATEQ4F, with revised thermodynamic data base and test cases for calculating speciation of major, trace, and redox elements in natural waters. US Geological Survey Open-File Report: US Geological Survey, 1991, 188.

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E. González, E. Forero (Eds) Bio-Nanotechnology for Sustainable Environmental Remediation and Energy Generation. ACCEFYN&NanoCiTed, Bogotá, 2016.

Arsenic in drinking water: Current situation and technological alternatives for removal Ma. Teresa Alarcón-Herrera1, Alejandra Martín-Domínguez2 Liliana Reynoso Cuevas1, M. Piña-Soberanis2 A. González-Herrera2.

A

rsenic pollution of natural and/or anthropogenic origin represents a global health challenge that affects millions of people around the world, especially in Latin America. This problem is particularly pernicious because regions polluted with arsenic remain mostly unperceived, the analytic identification of this metalloid in water is not obvious, and the adverse effects to human health are chronic and hard to directly associate. In affected communities, exposure to arsenic can be effectively mitigated by limiting the consumption of arsenic-laden water. Centro de Investigación en Materiales Avanzados, S.C. (CIMAV), Unidad Durango Victoria 147 nte, Centro Histórico, C.P. 34000, Durango, Dgo., México e-mail: [email protected] 2 Instituto Mexicano de Tecnología del Agua (IMTA). Paseo Cuauhnáhuac 8532, Col. Progreso, C.P. 62550, Jiutepec, Mor. e-mail: [email protected] 1

41

Bio-Nanotechnology

Introduction Arsenic is a common element found in the atmosphere, rocks, soil and water. It is moved to the environment through a combination of reactions that include natural processes and several by-products of human activity such as mining waste, fossil fuels, pesticides, herbicides, desiccants, wood preservatives, food cattle additives, semiconductors, pigments, among many others. People can be exposed to arsenic through inhalation and food or water ingestion. In certain areas of the world, the natural geology increases arsenic content in drinking water available to populations. Arsenic is highly toxic in its inorganic form; it is classified as a carcinogenic compound in the IA group (carcinogenic to humans) due to evidence it causes adverse health effects [1]. The long-term health effects of arsenic are the most worrying ones. These are mainly attributed to drinking arsenic-contaminated water, using it in food preparation, or ingesting food irrigated with it. Chronic toxicity produced by the accumulation of arsenic in the body results in skin lesions (e.g. hand and foot hyperkeratosis), myocarditis, diabetes, cardiovascular diseases, and damage to the nervous and respiratory systems. The permanent intake of contaminated water by arsenic causes the so-called “endemic regional chronic hydroarsenicism” (or HACRE by its initials in Spanish), which is commonplace in various parts of the world. Therefore, the presence of arsenic in surface waters (rivers, lakes, and reservoirs) and groundwater (aquifers) that can be used for human consumption represents a major health risk. To limit the adverse effects to exposed human populations, international institutions such as the World Health Organization (WHO) and others have established an arsenic concentration limit for drinking water of 10 µg/L. In México, the regulations currently 42

Arsenic in drinking water

establish that 25 µg/L is the maximum concentration limit in drinking water (modification to the Mexican official standard, NOM-127-SSA1-1994 [2]). This standard is under review and it is planned to decrease to 10 µg/L in the near future. Limiting the consumption of arsenic-contaminated water is an effective measure, which can mitigate the exposure of different affected communities. First, it is necessary to identify arsenic exposure by monitoring, measuring, and establishing appropriate remediation actions; these may include the use of alternative sources of groundwater or surface water and/or the use of novel technologies to remove arsenic from water. The following chapter focuses on the analysis of arsenic in drinking water and the alternative technologies used in Mexico for its remediation through water treatment.

Arsenic in water Arsenic can be found in water in both its organic and inorganic form, but the inorganic one is the most prevalent and it is considerably more toxic. In addition, arsenic occurs in soluble or particulate form. Depending on the local oxidation and reduction conditions, soluble inorganic arsenic subsists in two valence states. For example, arsenic in groundwater (under anoxic conditions) is in arsenite form (As+3) while, in surface water (under aerobic conditions), it is found in an arsenate or pentavalent form (As+5) [3]. Inorganic arsenic is present in water due to the natural dissolution of geologic deposits, industrial discharges, and atmospheric sediments. Because of this, arsenic concentration in the different environmental media has a high variability; however, water remains as its main dispersion pathway. 43

Bio-Nanotechnology

Arsenic distribution in superficial waters The arsenic content in rivers is relative low, generally less than 0.8 µg/L, but it can vary by orders of magnitude as a result of recharge (superficial and underground), geological composition of soil and rocks, weather, mining activity, and urban and industrial waste disposal. In semiarid regions, evaporation favors an increase in arsenic concentration in superficial waters, as well as an increase in water salinity and pH. In fact, this evaporative saturation has caused extremely high arsenic concentrations in places like the Loa River in the north of Chile (190–21,800 μg/L) [4]. In seawater, the average arsenic concentration is relatively low (1.5 to 4 µg/L). Estuaries have a variable arsenic concentration that result from continental waters, contribution of continental deposits, and local variations in salinity and redox gradients. The pluvial water of mining zones has a high arsenic content, usually ranging from 200 to 400 µg/L [5]. Arsenic contamination of geothermal origin in superficial waters and shallow aquifers has been usually reported in geothermal areas all over the world. In Mexico, high arsenic concentrations in geothermic zones has been associated to the Trans-Mexican volcanic belt. The documented cases are mainly in Los Azufres, Michoacan and Los Humeros, Puebla, where reported concentrations are estimated to be 800 µg/L [6].

Arsenic in underground waters Most aquifers with high arsenic concentrations have been linked to natural geochemical processes. One of the singularities with natural sources of arsenic in underground water is that there isn’t always a direct relationship between high arsenic content in water and high content of arsenic in the aquifer’s constituent materials. There is currently no geological/hydrogeological model for all the 44


identified incidents. Water with arsenic has been found in a wide variety of conditions, in oxidant and reductant environments, in over-exploded aquifers, in dry and humid zones or even in superficial and deeply confined aquifers. This variety of situations is defined because of the peculiarity of the circumstances and processes that occur in each one of these cases, i.e., the arsenic presence in each case is the consequence of the specific geochemical environment and hydrogeological conditions. Unlike anthropogenic pollution, that generates an affection of local character, natural arsenic pollution can affect vast areas. Arsenic concentration in underground waters is fundamentally controlled by water-rock interactions within the aquifer, and concentration values can greatly differ from one environment to other [7]. Taken together, these features tied to arsenic content in aquifers that could be useful for human consumption constitutes a significant health risk.

Worldwide situation Naturally, high arsenic concentrations affect wide areas all over the world. Aside from the problem of arsenic presence in drinking water, underground water with high arsenic levels is commonly used to irrigate diverse crops. This practice has caused the accumulation of arsenic in soils and therefore increased the transfer of metalloids into the food chain [8]. Different studies have identified many areas that contain underground waters with an arsenic content higher than 50 µg/L. The problems most quoted in the literature are, in Asia, Bangladesh, India (West Bengal), Nepal, China, Taiwan, and Vietnam; in America, high arsenic concentrations have been reported in USA, Argentina, Chile, Peru, Bolivia, El Salvador, and Mexico; in Europe, Greece, Hungary, Romania, and Spain. Asia has been 45

Bio and Nanotechnology

the continent where most investigations have been carried out, especially in Taiwan and Bangladesh [9, 10].

Situation in Latin America In 2006, at least 4 million of people in Latin America were exposed to arsenic levels that could cause health alterations via the consumption of contaminated water. Rural populations are the most affected ones because of socioeconomic, cultural, and sanitary conditions. In 2008, the Iberoarsen network estimated that 14 million people were at risk of arsenic contamination [10, 11]. The situation has been getting more critical in recent years. The metalloid has been detected in new zones or in zones that had been previously deemed unlikely to be contaminated. Furthermore, population growth has increased both exposure and anthropogenic pollution (mainly by mining industry leachates). In Argentina, the origin of arsenic in groundwater is mainly due to the volcanic activity of the Andes. Arsenic concentration in groundwater varies within a wide range, from less than 10 µg/L to more than 5000 µg/L. The affected region covers roughly 1.7 million km2 and is considered one of the largest in the world; it includes the provinces of Córdoba, La Pampa, Santiago del Estero, San Luis, Santa Fe, Buenos Aires, Mendoza, San Juan, Chaco, Formosa, Salta, Jujuy and Tucuman [12]. In this region, the rural population (> 1.2 million people) depends on groundwater for their water consumption and for agricultural activities. Arsenic concentration in rural areas at the provinces of Santa Fe and La Pampa exceed 67 µg/L. In Santiago del Estero, which is located in the northeast of the province, arsenic levels in groundwater range from 400 to 600 µg/L. One of the areas most affected by high groundwater arsenic levels is the southeast of the province of Córdoba. The most affected departments are Union, San Justo, Marcos Juarez and Rio Cuarto, where groundwater arsenic concentration ranges from 10 to 4550 µg/L [13, 14]. 46


The north of Chile is reported as the most affected area and is located largely in the Atacama Desert and part of the Andean mountain range. Reported arsenic concentration in the water range between of 200 to 900 µg/L [4, 9, 10]. In Antofagasta and Calama communities, the population was exposed to high arsenic concentrations in drinking water, until treatment plants for arsenic removal from water were installed in the 70´s . later seawater desalination plants were operating in northern Chile to provide drinking water [10]. In Bolivia, the Lake Poopo basin and the city of Oruro are affected by natural and anthropogenic arsenic in different environments like dust in the air (atmosphere) soils (e.g. in Oruro, where toxic effects were observed in the population) and in surface waters with concentrations ranging from 10 to 2460 µg/L in places devoid of anthropogenic activity. In places with mining activities or environmental liabilities, concentrations range from 600 to 11,140 µg/L [15, 16]. The rivers in the southeastern part of the basin of Lake Poopo have concentrations up to 87 µg/L while the hot springs exhibit concentrations of 65 µg/L. The arsenic content in the south and west of Lake Poopo in water samples from wells (the main source of water for human consumption) have arsenic concentrations up to 299 µg/L. Communities north and northeast of Lake Poopo have reported concentrations up to 964 µg/L [16,17]. An analysis of the water supply wells for human consumption indicates that 90% of this exhibit an arsenic concentration that exceeds the 10 µg/L concentration recommended for human consumption by World Health Organization. In Paraguay, the reported concentrations are greater than 50 µg/L in the groundwater of the Guarani aquifer, Formación Misiones, and high Paraná [10, 11]. In Peru, arsenic contamination reported in the area of Llo is of 47

Bio-Nanotechnology

natural and anthropogenic origin in the surface and groundwater basins of the Andean Region. Arsenic levels have been determined to range between 200 and 400 µg/L in the basins of the Locumba and Rimac rivers [10, 11,18]. In Brazil, arsenic contamination is mainly anthropogenic and is derived from industrial processes related to gold, lead, and zinc mining in the area of Minas of Gerais, Valle Riviera, and the Amazon region [9, 11]. In Uruguay, most water sources available for human supply are from superficial waters. Therefore, the situation is not critical, and concentrations of arsenic in the Mercedes aquifer are on the order of 10-58 µg/L [10, 19]. In Honduras, in the region of Llopango area of Salvador, department of Santa Ana, and the valley of Syria, anthropogenic pollution results from the exploitation of gold and silver ores; in these regions, arsenic concentrations in water are about 50-750 µg/L [10, 11]. In Nicaragua, the volcanic formations in some areas of the country induce the natural contamination of groundwater with arsenic. This occurs in hydrothermally altered and mineralized structures that are primarily the source of arsenic and that are located in tectonic lineaments parallel to Graben of Nicaragua. Natural arsenic contamination in water for human consumption has been identified in the northwestern region (Villanueva, Santa Rosa del Peñon), North Central (Madriz, Nueva Segovia), central (Valle Sebaco) and the Department of Chontales (La Libertad, town Kimuna) [19, 20, 21]. In rural communities like Zapote (Sebaco Valley), Santa Rosa del Peñon, Cerro Mina de Agua, Kimura, Ciudad Dario, San Isidro and Las Pilas, exhibit arsenic concentrations in water up to 1200 µg/L [19, 20]. 48


In Costa Rica, arsenic concentrations vary between 10 and 200 μg/L in surface waters and aquifers, and these are attributed to active geothermalism in contact with volcanic muds [10, 19] In the Dominican Republic, studies report that arsenic concentrations in the surface water of the Magauca river basin and the Margajita streamare As+3 inorganic > As+3 organic > As+5 inorganic > As+5 organic > arsenical compounds 51

Bio-Nanotechnology

and elemental arsenic [35]. Organic arsenic compounds are less harmful to health and are mainly found in seafoods. The signs and symptoms associated with high levels of prolonged exposure to inorganic arsenic differ highly amongst individuals, population groups, and geographical areas. There isn’t currently a method to distinguish between cases of cancer caused specifically by arsenic and those induced by other factors; a reliable estimate of the magnitude of the problem worldwide is therefore lacking. Early symptoms of prolonged exposure to high levels of inorganic arsenic are found in the skin, these include skin lesions, calluses on the palms of hands and soles of the feet (hyperkeratosis), hyperpigmentation, and hypopigmentation. Non-cancerous effects have been reported, these include damage to the cardiovascular system, kidney and liver disorders, peripheral neuropathies and developmental encephalopathies, and disruption of the endocrine system related to the development of diabetes [36, 37]. In relation to its carcinogenic effects, a relationship between arsenic in water and an increased presence and mortality from cancers of the bladder, lung, kidney and liver cancer in the exposed population have been identified. Air and occupational exposure to arsenic has been related to the development of bronchogenic cancer.

Prevention and Control The most important action to be undertaken by communities affected by arsenic pollution is to prevent a prolonged exposure by implementing a secure supply of drinking water for both consumption and food preparation. It is possible to mitigate human exposure by limiting the consumption of arsenic-contaminated water in different affected 52


communities by first identifying its presence through measurement and then establishing the appropriate mitigation measures, e.g. the use of alternative sources of groundwater or surface water and/or the use of diverse technologies for removing arsenic from water.

Arsenic removal technologies There is a wide range of adequate technologies to remove arsenic from water, either intradomiciliary or by centralized systems. The following technologies are among the most studied: coagulationfiltration, coagulation-flocculation-sedimentation-filtration, lime softening, ion exchange, reverse osmosis, nanofiltration, coagulation-microfiltration, coagulation-ultrafiltration, electrodialysis, capacitive deionization, adsorption on activated alumina or minerals containing iron, adsorption on nanomaterials, electrocoagulation, and distillation. Most of these processes have already been evaluated and validated at both laboratory and field -pilot and full-scale level [38, 39, 40]. The scope of each technology is influenced by diverse variables. For example, the volume of water to be treated, the kind of contaminants present in the water, the availability of trained staff, the feasibility of installing a centralized system with respect to the size and density of urban areas, the user’s economic capacity, the area in which to install the system, and costs associated with operation and maintenance. Additionally, the type and amount of waste generated during these processes is another important element to consider in the selection of the technology. The residue should be handled and properly disposed in order to avoid contamination risks caused mainly by leachates. The United States, Argentina, Chile, and recently Mexico account for most of the use, in America, of conventional processes. Coagulation technology has been used since 1970 in Chile. 53


The membrane processes (Reverse Osmosis, Nano, Micro and Ultrafiltration) and recently direct filtration are the most widely systems used both in Argentina (since 1970) and Mexico (since 1990) [41, 42, 43].

Intradomiciliary Systems An intradomiciliary system is the kind of system that is installed inside a housing unit where residents have access to treated water. These systems are also known as equipment used at family level. In general, adsorption based processes are the most readily applicable as intradomiciliary systems, because the user need only pass the water through the filters. However, the main drawback is that the users are responsible for the proper use of the system and its efficiency. There is also the uncertainty of the lifetime of the adsorbent medium, which depends directly on the amount of water passing through it, which in turn depends on the number of inhabitants in each home and their consumption habits. The duration of a domiciliary filter is calculated relative to a population mean and an average of water consumption (2L/ person/day). The population that does not conform to these values will be at risk; on the other hand, the volume of consumption does not take into account the water used to prepare food. When this solution is applied, it must ensure an adequate mechanism to change filters when they are exhausted, and make a proper disposal of arsenic-saturated mediums. Another drawback of intradomiciliary solutions that use filters is that, normally, these systems have a tendency to be “fouled” over time. Consequently, their use is discontinued as they become impractical because of the water flow decline. The bacteriological aspect of these filters often is not considered, so the bacteria growth inside can affect water quality if appropriate disinfection 54


procedures are not taken. To date most filters have been distributed in the Mexican state of Durango at the household level in urban zones; however, this is a temporary and costly measure that does not solve the problem and only masks it.

Centralized systems Centralized systems also have advantages and disadvantages. On one hand, at drinking-water treatment plants, the quality of water distributed to residents and the waste management are easier and more reliable; however, any failure affects all treated water. Direct filtration (DF) is within the cheapest and easiest methods to operate, basically, an iron-based coagulant and chlorine are added to the pipe at the inlet to the filters. Filtration is carried out in a dual granular medium constituted by sand-anthracite with filtration rates below 7m3/m2/h, and for arsenic concentrations equal or less than 100 µg/L. The advantage of this system, at least in Mexico, is that all the items for construction and operation can be acquired within the country. The process of coagulation-flocculation-sedimentation-filtration is one of the variants of direct filtration. This process is used when the concentration of arsenic is greater than 100 µg/L [44, 45]. In these cases, the clarification step is required to avoid excessive filter backwash caused by the dosage of coagulant needed to remove the arsenic. This process is the most used in Mexico (Figure 1) for the treatment of drinking water, but it is also an efficient method for removing arsenic [46]. The chemical coagulation stage of both systems can be replaced by electrochemical coagulation. This process, called electrocoagulation, allows for the production of the coagulant of interest (Fe2+ in this case) by applying an electric current to a metallic electrode for its oxidation [47, 48, 49]. The advantage of 55


Figure 1. Arsenic removal plants installed in different communities in the north of Mexico, using the direct filtration process. (Source Mexican Institute of Water Technology, IMTA).

this process is that the dissolved solids, combined with the cation used for the chemical coagulation, are not added to the water. Thus reducing the amount of sludge produced as waste. The disadvantage in this case is a higher cost than other processes and higher operational complexity and maintenance [50]. The filtration step with sand-anthracite may be replaced by micro or ultra-filtration membranes, both systems allow for better control of the quality of treated water, but they represent higher investment costs than the granular material filters. They also have the disadvantage that the consumables are imported. Additionally, the highly qualified human capital required to operate these technologies is not always easy to find and keep, by water utilities companies in Mexico. Adsorption on iron-based media, either granular or nanoparticles, is a very efficient way to remove arsenic from water; nevertheless, the main drawback is the same as the membrane processes: these systems have only been tested on a laboratory scale and the experience at pilot level is minimal [51]. Another disadvantage of adsorption processes for arsenic removal 56


is the adsorption media. When completely saturated, the media has to be replaced and properly disposed, which also involves additional costs. Often, when the water utility company does not have the financial resources to replace the adsorbent media for new ones, the adsorption system is used as FD by means of coagulant addition. Although in both systems filters are used, the problem of converting an adsorption system to FD is that the filters’ design is not exactly the same and therefore, the FD efficiency will be severely affected because it has to adjust to adsorption columns. The main problem is the different filtration rates and backwash required for both systems. Regarding the costs and the simplicity of operation of the aforementioned arsenic removal systems, whether membrane, electrochemical, or physical, they cannot compete with direct filtration. This is essentially because the coagulant dose required is low (20-60 mg/L Fe for every mg/L As) and energy consumption can be reduced if gravity filters are used instead of pressure filters. In addition, even in pressure filters pumping requirements are no greater than 1.5 kg/cm2 during filtration. Reverse osmosis, nanofiltration, and electrodialysis have an additional disadvantage: the disposal of their water reject can only be disposed in the sewer system according with NOM-002SEMARNAT-1996, which sets the maximum permissible limits of pollutants in wastewater discharges to urban or municipal sewage systems, evaporation lagoons built for this purpose, or discharging them into the sea through submarine issuers. The latter is limited to the drinking water treatment plants available in coastal areas. The processes involving coagulation and/or precipitation have the advantage that the waste generated in the filters’ backwash and the sludge in the clarifiers can be dehydrated. This form allows for the disposal of dry waste; however, waste needs to be disposed of on land free of garbage leachate in order to prevent decrease the 57

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waste’s pH. This type of waste is not considered hazardous as long as no reductive conditions promote the dissolution of arsenic in water streams crossing the waste deposits. In Mexico, in the state of Chihuahua, there have been about 600 reverse osmosis plant installations in rural areas to date, with a capacity to remove arsenic of 1,000 to 5,000 liters per day. This is for water that is used only for human consumption [52]. These are small-scale centralized systems that work as commercial bottling plants, but in this case these are not for profit and their aim is to recover the operation costs. Their main drawback is that, more often than not, operators do not have sufficient technical training to detect when the water no longer satisfies the regulations and the indispensable change of membranes and/or filtration materials. Disinfection represents a financial and logistical problem that not all communities can solve. In this case, appropriate coordination with municipal or state governments is an indispensable factor. Full-scale systems installed at the wellhead in the Comarca Lagunera region, Mexico, to remove arsenic from supply sources with DF systems (Figure 1), were built with federal, state, and municipal resources and are being operated with a treated water cost ranging from $ 0.6-0.8/m3. This cost includes chemical reagents, electrical power for filter backwashing, staff and maintenance [52, 53, 54]. The municipal water utility currently controls these water treatment plants. This operating cost is the most competitive in Mexico. Moreover, as long as the technology needed to minimize the cost of systems of adsorption and/or membranes is not built in the country FD or conventional clarification (depending on the concentration of As) will continue to be the most viable processes for arsenic removal in countries like Mexico. On the other hand, the correct operation of drinking water treatment systems is one of the major problems 58


that must be overcome to guarantee the availability of arsenic free water for the people, as well as constant monitoring of water quality and waste disposal.

Conclusions Water contamination with arsenic is a serious problem that requires immediate attention and action. Unfortunately, new zones that implicate a potential risk to the population are being detected in Mexico with more frequency and these are mainly due to an overexploitation of aquifers. On the other hand, the little information that users have on this issue, tied to the insufficient monitoring coverage and the costs associated with water purification, are all factors that increase the exposure of the population to this contaminant. Education and community involvement are key factors to ensure effective results in any of the actions to be undertaken. It is necessary that at least the institutions responsible for water management in each locality understand the risks associated with exposure to high levels of arsenic, either by consumption of arsenicladen drinking water, through food cooked with contaminated water, or via food crops irrigated with contaminated water (e.g. rice). The technology exists to remove arsenic. In Mexico, Direct Filtration has proven to be a technically and economically viable process for removing arsenic from water. However, its applicability depends highly on the type of water to be taken in the specific community and it cannot be considered a single solution applicable to all communities. Decision makers should carry out studies to ensure that the selected option is the most convenient from a technical, economic, and social point of view. The main barrier for the mitigate of problems is the lack of 59

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information, education, and awareness of decision makers and water managers. Also, more research and collaboration between institutions and involved agencies is required before the best options for each specific situation in each community is found.

References [1] ATSDR, home page. http://www.atsdr.cdc.gov/toxprofiles/tp2.pdf (accessed Jan 10, 2016). [2] NOM 127-SSA1-1994, Norma Oficial Mexicana, “Salud Ambiental, Agua Para Uso Y Consumo Humano-Límites Permisibles De Calidad Y Tratamientos A Que Debe Someterse El Agua para su Potabilización. home page. http://www.salud. gob.mx/unidades/cdi/nom/127ssa14.html [3] Arsenic Treatment Technology Evaluation Handbook for Small Systems; United States Environmental Protection Agency, EPA Office of Water, (4606M) 2003, EPA 816-R-03-014. [4] Cáceres, D.D.; Pino P.; Montesinos, N.; Atalah, E.; Amigo, H.; Loomis, D. Exposure to inorganic arsenic in drinking water and total urinary arsenic concentration in a Chilean population. Environ. Res. 2005, 98(2), 151-159. [5] Galindo, G.; Fernández Turiel, J. L.; Parada, M.A.; Torrente, D.G. II Seminario Hispano Latinoamericano Sobre Temas Actuales de Hidrología Subterránea. IV Congreso Hidrogeológico Argentino Río Cuarto, 25 al 28 de Octubre de 2005. http://digital.csic.es/bitstream/10261/4019/1/Galindo_et_al-Arsenico-2005.pdf (accessed Jan 24, 2016). [6] Carrillo-Chávez, A.; Morton-Bermea, O.; González-Partida, E.; RivasSolórzano, H.; Oesler, G.; García-Meza, V.; Hernández, E.; Morales, P.; Cienfuegos, E. Environmental geochemistry of the Guanajuato Mining District, Mexico. Ore Geol. Rev. 2003, 23, 277-297. [7] Smedley, P. L.; Kinniburgh, D.G. A review of the source, behavior and distribution of arsenic in natural waters. Appl. Geochem. 2002, 17, 517-568. [8] Zhu, Y.G.; Williams, P.N.; Meharg, A.A. Exposure to inorganic arsenic from rice: A global health issue? Environ. Pollut. 2008, 154, 169-171.

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[9] Bundschuh, J.; Litter, M. I.; Parvezg, F.; Román-Rossh, G.; Nicolli, H.B.; Jeanc, J.S.; C.W. Liu, C.W.; Lopez, D.; Armienta M.A.; Guilhermem, L.R.G.; GomezCuevas, A.; Cornejo, L.; Cumbal L.;Toujaguez R. One century of arsenic exposure in Latin America: A review of history and occurrence from 14 countries. Science of The Total Environment. 2012, 429, 2–35. [10] Bundschuh,J.; Pérez Carrera, A.; Litter, M. IBEROARSEN, Distribución del arsénico en las regiones Ibérica e Iberoamericana. 2008, CYTED Argentina, ISBN 13978-84-96023-61-1. [11] Morgada, M.E.; Mateu, M.; Bundschuh J.; Litter, M.I. Arsenic in the Iberoamerican region. The IBEROARSEN Network and a possible economic solution for arsenic removal in isolated rural zones. e - T e r r a. [Online] 2008. http://e-terra.geopor.pt , (accessed May 3, 2016). [12] Smedley P.L.; Kinniburgh, D.G.; Macdonald, D.M.J. ; Nicolli, H.B.; Barros, A.J. ; Tullioc, J.O. ; Pearced, J.M. ; Alonso M.S. Arsenic associations in sediments from the loess aquifer of La Pampa, Argentina. Applied Geochemistry. 2005, 20, 5, 989–1016. [13] Pérez Carrera, A.; Fernández Cirelli, A. Arsenic and fluoride levels in water for dairy cattle (Province of Córdoba, Argentina). InVet. [Online] 2004, 6(1), 5159. ISSN (Online):1668-3498. http://www.fvet.uba.ar/publicaciones/archivos/ ant/perezcarrera.pdf (accessed April 20, 2016). [14] Pérez Carrera, A.; Fernández Cirelli, A. Problemática del arsénico en la llanura sudeste de la provincia de Córdoba: Biotransferencia a leche bovina. InVet [Online] 2007, 9(1), 123-135. ISSN (Online): 1668-3498. http://www.scielo.org. ar/pdf/invet/v9n1/v9n1a13.pdf (accessed May 7, 2016). [15] Ormachea Munoz, M.; Quintanilla Aguirre, J. Distribution of geogenic arsenic in superficial and underground water in central Bolivian highlands. Rev. boliv. quim. [Online] 2014, 31(2), 54-60. ISSN: 0250-5460/ 2078-3949. http:// www.scielo.org.bo/pdf/rbq/v31n2/v31n2_a04.pdf (accessed May 7, 2016). [16]Van Den Bergh, K.; Du Laing, G.; Montoya, J.C.; De Deckere, E.; Tack, F.M. Arsenic in drinking water wells on the Bolivian high plain: Field monitoring and effect of salinity on removal efficiency of iron-oxides-containing filters. J. Environ. Sci. Health A Tox. Hazard Subst. Environ. Eng. 2010, 45(13), 1741-1749. [17] Ormachea Muñoz, M.; Garcia, J.L.; Bhattacharya, A.P.; Sracek, O.; Garcia

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Moreno, M.E.; Kohfahl, C.; Quintanilla Aguirre, J.; Diaz,J.H.; Bundschuh, J. Geochemistry of naturally occurring arsenic in groundwater and surface-water in the southern part of the Poopó Lake basin, Bolivian Altiplano. Groundwater for Sustainable Development, 2016, DOI:10.1016/j.gsd.2016.04.002. [18] Reuer, M. K., Bower, N. W., Koball, J. H., Hinostroza, E., De la Torre Marcas, M. E., Surichaqui, J. A. H., & Echevarria, S. Lead, Arsenic, and Cadmium Contamination and its Impact on Children’s Health in La Oroya, Peru. International Scholarly Research Network, 2012, [Online]: ISRN Public Health. [19] Bundschuh, J.; Bhattacharya. Arsenic in Geosphere and Human Diseases; 2010 Taylor and Francis Group, London, ISBN978-0-415-57898-1. [20] Cruz, A.C.; Fomsgaard, I.S.; Lacayo, J. Lead, arsenic, cadmium and copper in Lake Asososca, Nicaragua. Sci. Total Environ. 1994, 155(3), 229-236. [21] McClintock, T.R.; Chen, Y.; Bundschuh, J.; Oliver, J.T.; Navoni, J.; Olmos, V.; Villaamil Lepori, E.; Ahsan, H.; Parvez, F. Arsenic exposure in Latin America: Biomarkers, risk assessments and related health effects. Sci. Total Environ. 2012, 429, 76-91. [22] Alonso, D.L.; Latorre, S.; Castillo, E.; Brandão, P.F.B. Environmental occurrence of arsenic in Colombia: A review. Environ. Pollut. 2014, 186, 272–281. [23] Alfaro, R.; García, E.; Montenegro, O. Niveles de contaminación de mercurio, cadmio, arsénico y plomo en subsistemas de producción de la cuenca baja del Río Bogotá. Rev. U.D.C.A Act. Div. Cient. 2002, 4(2), 66-71. [24] Miranda, D.; Carranza, C.; Rojas. C.A.; Jerez, C.M.; Fischer, G.; Zurita, J. Accumulation of heavy metals in soil and plants of four vegetable crops irrigated with water of Bogota river. [Online] Rev. Colomb. Cienc. Hortíc. 2008, 2(2), 180-191. http://www.soccolhort.com/revista/pdf/magazin/Vol2/vol.2%20no.2/ Vol.2.No.2.Art.5.pdf . [25] Gleason, S.V. Riesgo sanitario ambiental por la presencia de arsénico y fluoruros en los acuíferos de México. [Online] 2002. www.bvsde.paho.org/ bvsaidis/mexico13/104.pdf (accessed April 20, 2016). [26] Ortega-Guerrero, M. A. Presencia, distribución, hidrogeoquímica y origen de arsénico, fluoruro y otros elementos traza disueltos en agua subterránea, a escala de cuenca hidrológica tributaria de Lerma-Chapala, México. Rev. mex. cienc. geol. [Online]. 2009, 26(1), 143-161. ISSN (Online): 2007-2902. http://

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[35] Castro de Esparza, M. L. Remoción del arsénico en el agua de bebida y biorremediación de suelos. Natural arsenic in groundwaters of Latin America, International Congress, Mexico City. CEPIS/SDE/OPS. HDT - No 96. ISSN: 1018-5119; March 2005. [36] Sergio Zarazúa, Rosalva Ríos, Juan Manuel Delgado, Martha E. Santoyo, Deogracias Ortiz-Pérez, María E. Jiménez-Capdeville. Decreased arginine methylation and myelin alterations in arsenic exposed rats. NeuroToxicology, 2010, 31, 1, 94-100.

[37] OMS Home Page. http://www.who.int/mediacentre/factsheets/fs372/es/ (Accessed Jan 20, 2016).

[38] Pérez-Castrejón, S.; Rivera-Huerta, M.L.; Martín-Domínguez, A.; GeloverSantiago, S.L.; Gómez-Rojas, A.; Hernández-Yáñez, C. Electrocoagulación a escala piloto para la remoción de arsénico en agua para consumo humano. XXXII Congreso Interamericano de Ingeniería Sanitaria y Ambiental, AIDIS. Punta Cana, Bávaro, República Dominicana. 7 al 11 de noviembre del 2010. Pp. 1-8. [39] Mundo-Ávila, E.; Martín-Domínguez, A.; Calderón-Mólgora. C. Remoción de arsénico en agua para consumo humano mediante el proceso de electrocoagulación-microfiltración. IX Congreso Regional para Norteamérica y el Caribe “Retos ambientales y oportunidades en Norteamérica y el Caribe. San Juan de Puerto Rico. 18 de noviembre de 2011. [40] Calderón-Mólgora, C.G.; Arroyo-Martínez, P.; Cruz-Gutiérrez, F.V.; Garrido-Hoyos, S.E.; Gelover-Santiago, S.; López-Corzo, R.; Martín-Domínguez, A.; Pérez-Castrejón, S.; Quezada-Jiménez, M.L.; Rivera-Huerta, M.L.; SeguraBeltrán, N. Evaluación técnico-económica de cinco tecnologías para remoción de arsénico. Instituto Mexicano de Tecnología del Agua. Proyecto TC 0815.3. 2010. [41] Piña-Soberanis, M.; Calderón-Mólgora, C.G.; González, A.; MartínDomínguez, A. Asistencia técnica para la licitación e instalación de plantas potabilizadoras para remoción de arsénico en Gómez Palacio, Durango”, Informe final de proyecto IMTA IMTA TC-1124.3. 2011. 42] González, A.; Calderón-Mólgora, C.G.; Montellano, L.; Piña-Soberanis, M. Estudio y asesoría técnica para la licitación e instalación de plantas potabilizadoras a pie de pozo para remoción de arsénico en Torreón, Coahuila; segunda etapa, Informe final de proyecto IMTA TC-1243.3, Convenio de Colaboración SIMAS Torreón-IMTA. 2012.

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[43] González, A.; Calderón-Mólgora, C.G.; Piña-Soberanis, M.; OrdóñezMartínez, A.; Flores, A.; Martín-Domínguez, A.; Calleja, D. Elaboración de los proyectos ejecutivos de las plantas potabilizadoras para remoción de arsénico de los pozos 35, 16-R y 50-R del municipio de Torreón, Coahuila”, informe final de proyecto IMTA TC 1327.3 del Contrato de Prestación de Servicio IMTAConsultora, Prestadora y Servicios de la Laguna, S. A. de C. V. 2013. [44] Rivera-Huerta, M.L.; Pérez-Castrejón, S.; Martín-Domínguez, A.; GeloverSantiago, S.; Gómez-Rojas, A.; Hernández-Yánez, C. Análisis técnico y económico del proceso de coagulación con cloruro férrico para remover arsénico de agua subterránea. Revista AIDIS de Ingeniería y Ciencias Ambientales: Investigación, Desarrollo y Práctica. 2011, 4(1), 46-56. ISSN: 0718-378X. [45] Rivera-Huerta, M.L.; Cortés-Muñoz, J.E.; Piña-Soberanis, M.; MartínDomínguez, A. Remoción de Hierro y Arsénico de Agua para Consumo Humano Mediante Precipitación y Adsorción en Zimapán, Hgo. México. XXVII Congreso Interamericano de Ingeniería Sanitaria y Ambiental. Porto Alegre Río Grande Do Sul Brasil, 9 pág., 3-8 diciembre 2000. [46] CONAGUA 2014. Situación del Subsector Agua Potable, Drenaje y Saneamiento Edición 2014. Noviembre de 2014. http://www.conagua.gob.mx. [47] Martín-Domínguez, A.; Rivera-Huerta, M.L.; Piña-Soberanis, M.; PérezCastrejón, S. Reactor a flujo pistón para remover arsénico por electrocoagulación. Publicación de la Asociación Argentina de Ingeniería Sanitaria y Ciencias del Ambiente- AIDIS Argentina, Edición No. 94 de Septiembre/Octubre. 2007, 3946. ISSN: 0328-2937. [48] Martín-Domínguez, A.; Rivera-Huerta, M.L.; Piña-Soberanis, M.; PérezCastrejón, S. Incidencia del gradiente de velocidad en la eficiencia de la electrocoagulación para remover arsénico en un reactor a flujo pistón. Interciencia. 2008, 33(7), 496-502, ISSN: 0378-1844. [49] Rivera-Huerta, M.L.; Martín-Domínguez, A.; Piña-Soberanis, M.; PérezCastrejón, S.; García-Espinosa, J.E. Optimización de un floculador acoplado a un reactor de electrocoagulación a flujo pistón para remoción de arsénico. XXXI Congreso Interamericano de Ingeniería Sanitaria y Ambiental, 1 Foro Interamericano sobre Servicios de Agua y Saneamiento. Santiago de Chile. 8 pp, 12 al 15 de octubre de 2008. [50] Pérez-Castrejón, S.; Rivera-Huerta, M.L.; Martín-Domínguez, A.; GeloverSantiago, S.L.; Piña-Soberanis, M.; Gómez-Rojas, A.; Hernández-Yáñez, C.;

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Cortés-Muñoz, J.E. Comparación Técnico-Económica entre coagulación química y electrocoagulación para remover arsénico del agua. Tecnología y Ciencias del Agua. 2012, III, 5-22. ISSN: 2007-2422. [51] Qu, X.; Alvarez, P.J.J.; Li, Q. Applications of nanotechnology in water and wastewater treatment. Water Research. 2013, 47, 12, 3931-3946. [52] COFEPRIS Home Page. http://www.cofepris.gob.mx/Paginas/Temas%20 Interes/Programas%20y%20Proyectos/Agua/AguaCalidadFisicoquimica.aspx (accessed April 23, 2016). [53] González, A.; Montellano, L.; Calderón-Mólgora, C.G.; García, A.; FloresOcampo, A.; Islas, A.; Piña-Soberanis, M. Selección de trenes de tratamiento e ingeniería básica para potabilizar 26 fuentes de abastecimiento y elaboración de proyectos de seis plantas potabilizadoras en la Región Lagunera, estado de Coahuila de Zaragoza, Informe final de proyecto IMTA TC1408.3, 30-39. Convenio de Colaboración CONAGUA-IMTA SGAPDS-OCCCN-Rl-14-040-FCC. 2015. [54] González, A.; García, A.J.; Flores-Ocampo, A.; Islas, J.A.; Rodríguez, J.; Sánchez, H.; Piña-Soberanis, M. Ingeniería de detalle, implementación, puesta en marcha e inicio de operación de la infraestructura que permita mejorar el abasto y la calidad del agua en la Región Lagunera en el estado de Durango.”, Informe final de proyecto IMTA TC1522.3, Convenio de Colaboración CONAGUAIMTA SGAPDS-OCCCN-RL-15-06-FED-CC. 2015.

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Modelling of mercury transport, fate and transformation in continental surface water bodies A case study of the Mojana Region, Colombia Nelson Obregón1, Leonardo García2, Diana M. Muñoz1

S

cientific knowledge related to the dynamics of mercury in the environment and precisely in aquatic ecosystems has been focused in understanding the relationships and processes that control mercury transport, fate and transformation from the different sources of input into aquatic ecosystems to bioavailability processes and bioaccumulation in food chains. Diverse mathematical models of various characteristics have been developed with the purpose of simulating mercury species transport, fate and transformation processes in aquatic ecosystems as tools to approach mercury dynamics in these ecosystems using mathematical expressions, and to use the results to contribute to decision-making related to the management of mercury contamination problems, ecological hazards and human health effects. Geophysical Institute, Faculty of Engineering, Pontificia Universidad Javeriana Bogotá, Colombia e-mail: [email protected] 2 Basic Science Department, Universidad Jorge Tadeo Lozano, Bogotá, Colombia e-mail: [email protected] 1

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Introduction Mercury is a persistent environmental contaminant that does not degrade. However, it changes in shape and moves through various environmental compartments (air, water, biota, soil and sediment) [2]. This heavy metal is known for its toxicity and negative effects on human health and natural ecosystems [1]. Among the species of organic mercury, methylmercury (MeHg) is considered the most toxic species due to its bioaccumulative and biomagnification capabilities through food chains [4]. The most common form of methylmercury exposition for human beings is through the ingestion of contaminated species, mainly of fish species [5]. The risks to human health due to chronic exposition to methylmercury generate severe socioeconomic consequences for the population [2]. Consequently, in the last few years there has been an increase in the attention to environmental mercury contamination reflected in the global efforts to reduce anthropogenic emissions to the environment and also in research focused in mercury dynamics in the environment, specifically in the quantification of its concentration, mobilization and transformation [2][4][5]. The sources of mercury in the environment include natural sources (e.g., volcanic emissions, discharge from natural mineral sources, forest and soil burning), anthropogenic sources (e.g., gold mining, fossil fuel combustion, industrial waste) and re-emission sources [2]. Regarding aquatic ecosystems, these are commonly contaminated with mercury due to the direct discharge or release produced by anthropogenic activities into water bodies. The most common of these activities are mining and industrial waste discharges combined with indirect sources such as atmospheric deposition, surface run-off, and soil erosion among others [3] [10][13]. In water bodies mercury is subject to transport, fate and transformation processes. The latter process is performed through multiple biotic and abiotic transformations such as photochemical 68

Modelling mercury transport, fate and transformation

reactions and microbiological activity reductions among which mercury methylation processes are the most significant. So far, methylation processes are not completely known, however, there is scientific evidence which indicates that in water bodies, methylation is produced by aerobic biological activity, particularly in highly productive hydrosystems in ecological terms [35] [43] [48]. Scientific knowledge related to the dynamics of mercury in the environment and precisely in aquatic ecosystems has been focused in understanding the relationships and processes that control mercury transport, fate and transformation from the different sources of input into aquatic ecosystems to bioavailability processes and bioaccumulation in food chains [12]. Diverse mathematical models of various characteristics have been developed with the purpose of simulating mercury species transport, fate and transformation processes in aquatic ecosystems as tools to approach mercury dynamics in these ecosystems using mathematical expressions, and to use the results to contribute to decisionmaking related to the management of mercury contamination problems, ecological hazards and human health effects [10] [11]. For these reasons, modelling could be a cost-effective way to assess management actions, for instance: the estimation of mercury dynamics in time to determine the risks to the ecosystem and human health, the effectiveness of actions to remediate or passively decontaminate ecosystems, the scope of control and management measures to reduce the sources of polluting loads [10] [11] [17]. However, modelling mercury dynamics in aquatic ecosystems is considered a complex activity [11] [12], due to the amount of processes and factors that govern transport and fate processes (hydrodynamic and sediment transport) and also transformation processes (biogeochemical) to be considered in water bodies. Such complexity is mainly represented in technical, economic and timeconsuming efforts [10]. 69

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The Mojana region, located in northern Colombia, is a territory that is classified as a floodplain through the formation of a delta in the confluence of the Cauca, Magdalena and San Jorge Rivers [54]. Such region is known for its biodiversity and for the ecosystem services it provides to the local population, to the MagdalenaCauca basin and to humanity, especially through its vast system of wetlands which is the most representative ecosystem of the floodplain [54]. However, the sustainability of the natural system of this region is endangered due to mercury contamination, which mainly affects aquatic ecosystems and their ecosystem services. Fishing is one of these. A significant proportion of the local population depends on fishing as a source of food, and fishing has developed into an economic activity of regional significance [52]. Several studies such as Marrugo (2015) [55], Pinedo et al. (2015) [56] and Olivero and Johnson (2002) [4] have shown the critical level of mercury contamination in the region through measurements in different environmental compartments combined with the toxicology effects. The sources of mercury contamination in the Mojana region are directly related and mainly caused by the extensive gold mining activity developed in upstream regions in the basins, including the largest and most intense mining region of the Lower Cauca river basin located in Antioquia and Bolivar departments, which yields an annual average of more than 30 % of the national gold production [50]. As a large proportion of this mining activity is conducted through artisanal and small-scale gold mining processes, known for the use of precarious technology, the absence of knowledge or any regulations or standards and the indiscriminate use of mercury for the gold extraction mining activities which a significant portion it is released into the environment becoming in one of the largest anthropogenic source of polluting loads in the Magdalena – Cauca river basin. Due to transportation phenomena in the different environmental compartments, a considerable portion of these contaminants ends 70


up in the aquatic ecosystems of the Mojana region, where they can transform, bioaccumulate and biomagnify in food chains and be exposed to the local population through ecosystem services like fishing [13] [55] [56]. Regarding the mercury contamination problem in the aquatic ecosystems of the Mojana region exposed above, the implementation of a model of transport, fate and transformation of mercury species in aquatic ecosystems could be a valuable tool for managing these environmental and toxicology regional issues and could contribute to the knowledge of this phenomenon’s environmental dynamics. This would allow us to simulate potential contamination scenarios and the efficacy of control or remediation measures, and these results could in turn feed cost-benefit analyses of such measures. Nevertheless, when a modelling project of these characteristics is undertaken it is fundamental to guarantee aspects such as technical capacity, sustainability, computation requirements and institutional commitment, among other factors, due to the complexity of the application of such models [10][12].

Mercury dynamics in continental surface water bodies Mercury is a metal that is naturally found in the environment, however, anthropogenic activities that have been in development for the last century or so have produced a significant increase in the quantity and distribution of mercury in the atmosphere. Such activities discharge or release mercury directly into aquatic ecosystems (e.g., industrial discharges, mining waste) or indirectly, by discharging or releasing mercury to the atmosphere, to the soil or subsoil through activities like fossil fuel combustion, metal mining, industrial activities, and inadequate disposal of solid waste containing traces of mercury, among others [1]. The point or non-point polluting loads released and stored in 71

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environmental compartments (i.e., air, vegetation, soil or subsoil) can be transported to surface aquatic ecosystems by natural processes like wet or dry atmospheric deposition of mercury, rain, surface and sub-surface run-off, soil erosion processes and the transport of sediment loads and particulate matter linked to mercury species transported into water bodies [2]. The remobilization of mercury is another indirect input source of this metal into aquatic ecosystems. This process encompasses the liberation of mercury species that were previously deposited or accumulated in several environmental compartments and released by human activities such as agriculture, reforestation, mining, dam construction, and sediment dredging or by natural physical biotic processes (hydrological, climatological, geological and biological among others) [3]. In surface water bodies, mercury is subject to various physical, chemical, biological and geological dynamics that will determine its transport, fate and transformation in the aquatic ecosystem. Among such processes, mercury transport and fate are significant due to the hydrodynamic effects of the water body, which is what happens with the advective-dispersive processes that lead the transport of dissolved mercury species and of those linked to suspended matter in the water body. This linking process between mercury species and solid particulate matter is due to the high capacity of mercury to be linked to solid organic or inorganic material, and it is highly significant in mercury dynamics of the aquatic environments as it affects its transport, which is why sediment transport processes (suspended solids, sedimentation and resuspension) are fundamental for the knowledge of the fate of mercury species, and their potential transformation or inactivation processes [2] [4]. After the stages of transport and fate in the aquatic ecosystem, these mercury species are stored in the different compartments of 72


the water body and subject to transformation processes through biogeochemical processes that will determine mercury speciation in the water, the bioavailability of mercury species, and also mass exchanges between the phases of the water body (bottom sediments and sediment -water column interface) related to diffusive mass exchange, biodispersion and hyporheic flow [5]. The results of several studies conducted by Gilmour et al. (1992) [6], Benoit et al. (1999) [7], Monperrus et al. (2007) [8] and Drott et al. (2008) [9], stated that mercury transformation processes in the different compartments of the water body are directly related to the microbiological activity of certain types of species found in the water column and the bottom sediments. These biogeochemical transform reactions convert existing mercury into reactive mercury (HgII), elemental mercury (Hg0) and methylmercury (MeHg), and their activity and speciation processes are highly dependent on environmental conditions. These species of mercury interact with the hydrobiological species present in the aquatic ecosystem and produce bioaccumulation processes in the hydrobiological species and biomagnification processes in the food chain. Due to human beings are exposed to contact and ingestion of these species, they represent a public health threat of the communities based in the mercury toxicity properties [4]. Figure 1 outlines the main specific processes of mercury dynamics in surface water bodies.

Characteristics of surface water mercury models of transport, fate and transformation The mathematical modelling of mercury dynamics in the environment, particularly the models that allow the simulation of the transport, fate and transformation of mercury species in aquatic ecosystems, is a tool that we can use to approach mercury pollution problems in water bodies and determine the hazards that such contamination can represent for the ecological state 73

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Figure 1. Specific mercury processes in surface water bodies.

and for the exposed human population [10]. Moreover, these models become fundamental tools in the formulation of plans and actions for the mitigation and control of mercury contamination in aquatic ecosystems. Through the results produced from scenario simulations using these models, represented in mercury species concentrations obtained from the models in the different compartments of simulated water bodies, they provide essential information to guide decision-making processes aimed at managing environmental contamination and the toxicology threat on human health that mercury contamination poses [5]. One of the main applications of mathematical models of mercury transport, fate and transformation in aquatic ecosystems is the evaluation of the risks derived from the contamination of these ecosystems and the remediation possibilities [11]. This is a costeffective measure in terms of economics and resource investment 74


that results from the application of a model of these features. A simple and effective measure in many cases, as opposed to other actions related to contamination management such as investment in technology, education, and restrictions to ecosystem services among other aspects [12]. Between the various remediation measures for aquatic ecosystems with mercury pollution problems, it is possible to simulate as a modelling scenario the natural attenuation of the mercury contamination problem trough time. The results obtained from the simulation of this scenario in the mercury species dynamics models in water bodies could allow to the authorities, stakeholders or environmental managers to approach the time scale in which the water achieves its own self depuration plus the final fate of these mercury species as a natural reaction of the ecosystem. Therefore, the application of these models could be a fundamental tool for assessing the viability of remediation actions in mercurycontaminated water bodies [11] [13]. The application of these models to mercury contamination in aquatic ecosystems would also provide information through the construction of prospective scenarios where mercury discharge into the aquatic ecosystem is reduced or increased as an environmental control measure. From the simulation performance results of the effects from these contaminant-variation scenarios, the effectiveness of the measures could be assessed to a certain degree [14]. In the specific case of determining hazards to human health due to mercury contamination in a water body, the models that would allow the simulation of the dynamics of mercury species transformation and bioaccumulation in the water body ichthyic fauna can provide results on bioaccumulated mercury in hydrobiological species and so specify the threat of mercury contamination to the human 75

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population due to the exposition to contaminated ichthyic species [5]. For the detailed case of transformation models for mercury species that allow the simulation of methylation and demethylation processes combined with the influence of environmental variables in the biogeochemical kinetics that produce methylated species, the results serve as a tool to approach methylation processes, bioavailability of methylated species in the water column, and environmental effects in their transformation kinetics. This information becomes a fundamental tool for the evaluation of mitigation measures for mercury contamination, due to the implications of methylation processes on the exposition of these toxic substances to hydrobiological species and humans [11]. [10] The complexity level of the surface water mercury models of transport, fate and transformation depends on the degree of detail in terms of spatial and temporal dimensions, the morphological characteristics of the aquatic ecosystem to be simulated, and the state or exact destination of the mercury species (biotic community, water body, sedimentation, humans, among others), for these reasons it is important to measure these aspects before developing a modelling project of these characteristics [12]. Additionally, the complexity of the modelling process is also reflected in the large amount of required data about the aquatic ecosystem, such as input data, and for the calibration and validation stages of the model. The reliability of the model and thus the adequacy of the decisions made regarding actions to take in order to face the contamination problem based on the results obtained from the applied models depend on these conditions [15].

Review of surface water mercury models of transport, fate and transformation The development of these models started around the mid 1960s as a response to the evidence of the negative effects of mercury 76


for the environment and its toxic effects in humans [9]. This explains why the main objective of the scientific community in the development of these models to simulate mercury dynamics in surface water bodies is to use them as tools to manage mercury contamination [14]. Thus far, multiple mathematical models with diverse applications and complexity levels have been developed. In general terms it is possible to classify these models in the categories below according to their modelling goal, spatial dimension, and types of simulated processes [10]. a) Models based on the exchange of mercury concentrations between the water-air-sediment interface through the balance of mass among these compartments via simplified processes, i.e., without considering the transport processes of dissolved mercury fractions or mercury linked to particulate suspended matter in the water body [16][17]. b) Models in which mercury is a conservative substance transported in the water column and linked to particulate suspended matter and its fate depends on hydrodynamic processes and also on the transport of sediments in the hydrosystem [18]. c) Models based on the geochemical processes of the mercury cycle in surface water bodies [19] [20]. d) Models based on the bioaccumulation and biomagnification of mercury species in aquatic ecosystem species and their food chain [12]. e) Adapted transport and fate models that allow the simulation of physical processes such as hydrodynamics, sediment transport, and biogeochemical transformations of mercury, using spatial approximations of one-dimensional, 77

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two-dimensional or three-dimensional models in surface water bodies [21]. Using the classification above, in the following section we will expose the characteristics, scopes, processes and requirements of these models in terms of the simulated processes in surface water bodies, the spatial dimensions of the simulation of the models, as is the case in hydrodynamic processes and sediment transport, biogeochemical processes and bioaccumulation processes.

Simplified mass balance models in well – mixed surface water bodies (zero-dimensional) The main object of these models is to simulate with a simplified approach the dynamics of mercury in surface lentic hydrosystems, without considering the effects of hydrodynamic and sediment transport, and therefore aquatic systems are regarded as a wellmixed reactors or zero-dimensional systems. Some of these models divide the water bodies into multiple layers between which uniform and time-constant mass exchanges occur [22]. This mass balance approach for mercury modelling cycle is appropriate for small reservoirs, lakes and wetlands where organic and inorganic suspended solids are present and along with minimum variations in terms of inputs or outputs to and from the hydrosystem, polluting loads, and hydrodynamic behavior. Thus, this model is inadequate for highly dynamic aquatic systems such as lotic, large lentic water bodies and coastal areas, where hydrodynamic behavior and sediment transport significantly impact the transport and fate of mercury [23]. Additionally, these models take into account the balance of mercury species in the different compartments through the concept of partition coefficients (KD) that depend on environmental variables of the aquatic ecosystem such as temperature, pH, and 78


redox potential among others that indirectly influence the kinetics of mass exchange, which implies the need to determine such environmental variables as input data for this type of models [24]. Among the pioneer models of this type the one developed by Turner and Lindbergh in 1978 [25], it is best known for their simulation processes which describes the dynamics of mercury concentrations in a river-reservoir aquatic system located downstream from the discharge of a chemical product plant. However, this model does not consider the effects of fractions of mercury linking to suspended solids. The results of these models overestimated mercury concentrations downstream from the discharge point of polluting loads [10]. Years later, Fontaine (1984) [28] developed a complete-mix model on the water body, considering three fractions of mercury species (dissolved, particulate and linked to organic substrates) combined with reaction processes, mass exchange between the water column and the bottom sediments controlled by kinetics, sorption processes and geochemical transformations on mercury species [12]. The QWASI (Quantitative Water-Air Sediment Interaction) model developed by Mackay and Diamond on 1989 [27] has the capability of simulating the fate of mercury from the dry and wet atmospheric deposition into the water body, combined with fate processes in the water column, in a completely well-mixed hydrosystem in stationary conditions, as seen in the speciation between dissolved fractions of mercury and those linked to suspended particulate matter, the diffusive mass exchange between water column and sediments, re-suspension and sedimentation [11]. Another model known for its application in various cases that has the capacity of dynamically simulating mercury fate and the mercury cycle in water bodies is the MCM (Mercury Cycling 79

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Model), developed by Hudson et al. (1994) [28]. This model’s main feature is the segmentation of the water body in multiple compartments, such as lentic hydrosystem, the epilimnion, the hypolimnion, and the bottom sediments, along with four hydrobiological compartments including zooplankton, phytoplankton, and carnivore and non-carnivore fish species. Regarding mercury species, this model simulates the following mercury species: elemental, reactive and methylmercury. With respect to the simulation of the latter species, it has the capacity to simulate reaction rates for methylation and demethylation as a function of the concentration of reactive mercury, sulfate ion, dissolved organic matter and water column temperature [10]. The MCM has been used in different studies, such as the modelling in large lakes in the United States developed by Leonard et al. (1995) [29]and Knightes and Ambrose (2007) [30] where they applied the model in 91 lakes in Vermont and New Hampshire (US), and Kotnik et al. (2002) [31] in the Velenje River in Slovenia. The SERAFM model developed by Knigthes et al. (2008) [32] is a mercury cycle modelling software extensively spread and applied in multiple cases. This model is based on the balance of mercury species in surface water bodies considered completely mixed or zero-dimensional, along with the segmentation of the system in several layers. It possesses a user-friendly interface developed from databases [10]. The following are among the general features of the SERAFM: categorization of mercury species in elemental, reactive and methylmercury through five phases (solid, inert, phytoplankton, zooplankton, detritus, and dissolved organic matter) and three water body compartments (epilimnion, hypolimnion and bottom sediments), and an equilibrium condition regarding the simulation of sorption processes. The cases in which this model was applied include the research developed by Brown et al. (2007) [33] in artificial wetlands in Nevada (US) and the case developed by Canu et al. (2012) [34] in the Marano-Grado lagoon 80


(Italy).

Models based on hydrodynamics and sediment transport These models simulate the temporal and spatial distribution of mercury point or non-point polluting loads as inputs of hydrosystems, along with the role of sediment transport and fate due to its capacity to link with solid particles dissolved in the water column [15]. The hydrodynamic model and the sediment transport model can jointly simulate the behavior of mercury species released in water bodies that are transported by the water column, in dissolved fractions or linked to suspended solids. Consequently, these models can approach the temporal and spatial distribution of mercury species through the simulation of hydrological, hydraulic and sediment transport dynamics [12]. Normally, this type of hydrodynamic and sediment transport models are appended to a water-quality module through which the changes in the concentrations of the mercury species dissolved in the water column are simulated using the advection-dispersion equation [14]. In a subsequent stage, the results from these models (hydrodynamic, sediment transport and water-quality) are appended to specific modules (sub – models) that simulate the mass exchange and mercury species concentration flow exchanges between the water body phases such as bottom sediments, the water column and the atmosphere [12]. Among the models of this type one of the most important is the MERMC4 model developed by Henry et al. (1995) [35], which is appended as a quality module integrated into the WASP (Water Quality Simulation Program). It is an integrated dynamic quality model for diverse types of hydrosystems through which the hydrodynamic and sediment transport model is appended to the mercury cycle model based on the mathematical formulation of the MCM mentioned in the previous section of this chapter section [36]. 81

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In turn, the WASP software offers modules for modelling mercury transport and fate depending on the water body flow properties. For lotic water bodies there exists the RIVMOD model, which optimizes hydrodynamic processes and fit them to kinetic models for sorption and desorption processes in different compartments of the hydrosystem [36]. The most representative cases conducted using this modelling framework are the mercury transport model developed by Carroll et al. (2000) [37] in the Carson River (Nevada, US), and the model developed by Zagar et al. (2006) [21] in the Soca and Idrijca rivers in Slovenia.

Models based on biogeochemistry and the bioavailability of mercury species The models based on mercury transformation processes establish the mercury speciation in surface hydrosystems through biogeochemical process, specifically concerning methylation and demethylation mass exchanges between the water body phases (bottom sediments, water-sediment interface and water column), and the bioavailability of mercury species [12]. According to several authors [38] [39] [40], it is estimated that speciation between the different phases of the water body is directly related to the microbiological activity of certain types of particular species developed in the water column and in the accumulated bottom sediments. These biogeochemical reactions transform existing mercury into reactive mercury (HgII), elemental mercury (Hg0) and methylmercury (MeHg), and it depends on environmental conditions such as nutrient availability, traces of metals, temperature and radiation among others. Thus, it is necessary to consider these conditions and simulate them as well in mercury transformation processes. These are regularly simulated as a submodel that is appended to the general mercury transformation model [38]. 82


A commonly used methodology in models for the simulation of mercury transformation dynamics in the environment it is to assume equilibrium conditions and estimate the variations among the dominant mercury species in terms of reaction equilibrium coefficients which depend on multiple environmental factors [36] [12]. Condition that increases the complexity in the simulation dynamics of these coefficients [17]. However, existing models typically attempt to simplify the simulation of these processes assuming that the concentrations and reaction coefficients are not independent [12]. These mercury biogeochemical transformation models have been developed through time, at present these models allow the calculation of the equilibrium of mercury species in the dissolved, solid and gas states and sediment bound in the different compartments of an aquatic ecosystem (water column, bottom sediments, water-sediment interface, and water-atmosphere interface) [5]. The MINTEQ+ model, developed by Bhavsar et al. (2004) [41] stands out among this type of models due to its applicability and use in multiple cases. It is a detailed geochemical speciation model with chain reactions between the different water column phases and the transfer of mass between the environmental compartments (water, air and sediments) as a dynamic system linked to a hydrodynamic and sediment transport model. This model was applied for modelling heavy metals in various lakes in Canada [41]. Concerning the simulation of mercury methylation and demethylation processes attributed mainly to bacterial metabolic activity, these biological reactions have particular kinetics, which are typically slow and irreversible. The BIOTRANSPEC model developed by Gandhi et al. (2007) [42] it is another representative model that specifically simulates the mercury methylation dynamics through biogeochemical activity allowing the dynamic simulation of methylation and demethylation processes. There are also models that detailedly simulate the methylation 83

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processes by microbiological activity of specific bacteria groups such as iron-reducing and sulfate-reducing, and therefore it is necessary to detail redox and controlled kinetic microbiological degradation reactions in the different phases of the water body [12] [43]. The specific models for simulating this detailed microbiological kinetics are appended or linked to the results and processes of hydrodynamic and sediment transport models and the mass balances between phases of the water column. Among these models, the PHREEQC2, developed by Parkhurst and Appelo (1999) [44] is the most well-known. It is used for exploring the relationships between mercury methylation and the environmental conditions that affect microbiological processes. Additionally, it integrates into software that couple hydraulic onedimensional modules, mercury speciation and biogeochemical processes modules [12]. In the case of mercury models that simulate the interaction of mercury species with biotic species allowing the modelling of the ecological effects of mercury contamination [10] [5], in general terms, these models allow the simulation of mercury species amounts present in fauna and flora. The most common modelling strategies for this type of models are the use of multiplication factors according to the concentration of mercury species in the water bodies and to the morphologic characteristics of the biotic species, through which mercury concentrations are bioacumulated in the species and biomagnified in the aquatic ecosystem food chain [45]. There are also more detailed approaches to these ecological processes through the modelling of food chains and energy demands of the species through the assessment of their feeding habits, quantities ingested, and metabolism among others [46] [47] [48]. These detailed models allow approaching knowledge of mercury bioaccumulation and biomagnification in fish and other hydrobiological species. However, the application of these models is complex, due to the level of detail of the processes 84


and the variables needed data for their calibration.

Data requirements for surface water mercury models of transport, fate and transformation The data requirements for the implementation of a mercury transport, fate and transformation model depends on its level of complexity in terms of spatial and temporal scale, the complexity of the aquatic system to simulate, the state or destination of the mercury species on the environment to simulate (bottom sediments, food chain, methylation or demethylation) [5][12]. Table 1 includes a brief description of the basic data requirements for the development of surface water mercury models (input and calibration variables and parameters), from the model type classification according to their simulation goal, spatial dimension and processes simulated for each type listed in the previous section. Regarding the data requirements presented in Table 1 for each type of mercury transport, fate and transformation model, it is possible to identify how the information requirements for the application of the models increases as the complexity of the model in terms of detail degrees of mercury modelling processes, space and time scale, and type of water body among other aspects. These data requirements according to the type of model and their higher level of complexity can be directly connected to a higher level of investment in various resources such as economic, technical, human, space and time, and institutional resources among others [11 ][12][13]. For this reason, when a mercury dynamics modelling project in an aquatic ecosystem it is to be develop, it is recommended to perform a cost-benefit analysis of the resources required among those mentioned for the application of the models and thus find the appropriate model type according to the modelling object and available resources. 85

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Approaching the mercury contamination problem in continental surface water bodies in Colombia, pointing to the case of the Mojana region According to the National Study on Water conducted by IDEAM in 2014 [49], it is estimated that in Colombia approximately 205 tons of mercury are released into continental surface water bodies Table 1. Data requirements for surface water mercury models by model type. ID Model type

Model type

Data requirements •

Mercury species concentrations [THg], [MeHg] in air, water and sediment. Hydraulic information of flows, volumes or water levels of the water body for space and time coarse scale. Mercury atmospheric deposition on annual or monthly scale. Hydroclimatological variables: rainfall on monthly or annual scale, wind regimes. Features of specific polluting loads: location, average load of mercury species.

•

a)

Simulation of mercury concentration exchange between the water-airsediment interface

• • •

•

b)

Simulation of mercury in the water body as a conservative substance transported in the water column and linked to particulate suspended matter. Its destination depends on hydrodynamic processes and sediment transport.

Hydrodynamic transport: hydrosystems topology, roughness coefficient, flow discharges and flow velocity in control points. Sediment transport: grain-size distribution of suspended and bottom sediments, concentration of total suspended solid fractions (volatile and fixed fractions), dissolved solids and average density of the sediment particles. Total mercury concentration [THg] and methylmercury [MeHg] in the water body, specific and tributary streams polluting loads and, suspended sediment and bottom sediment loads.

•

•

86


•

•

c)

Simulation of geochemical processes of the mercury cycle in the water body.

• •

•

d)

Simulation of • bioaccumulation and biomagnification of mercury species in aquatic ecosystem species and their • food chain. • •

Concentration of total mercury species in the water column and its segmentations (surface, bottom, mid-depth), total mercury [THg], elemental mercury [Hg0] and methylmercury [MeHg] in suspended sediments loads and sediments accumulated and stabilized sediment in the bottom. Kinetics rates of biogeochemical transformation processes of mercury species (measured in lab or reference values) in the water column and bottom sediments. Presence or absence of sulfur-reducing and iron-reducing bacteria in the water column and bottom sediments. Environmental conditions that influence biogeochemical processes in the water column and sediments: temperature, solar radiation, organic mass percentage, dissolved oxygen concentration, nutrient concentrations, and sulfur and iron species concentrations. Mercury species concentrations in the water column and its segmentation phases (surface, bottom, mid-depth), total mercury [THg], elemental mercury [Hg0] and methylmercury [MeHg] suspended sediments and sediments accumulated in the upper bottom and in stabilized sediment in the lower bottom. Mercury species concentrations in plankton species of water bodies (periphyton, phytoplankton and zooplankton). Mercury species concentrations in algae and aquatic vegetation tissues. Mercury species concentrations in carnivore and non-carnivore ichthyic species tissue and organs. Mercury species concentrations in amphibious or terrestrial fauna species which the main food source are the aquatic fauna.

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e)

Simulation of mercury transport and fate that allow the simulation of physical processes such as hydrodynamics, sediment transport and mercury biogeochemical transformations in water bodies using space and time approximations from one-dimensional, twodimensional and threedimensional models.

•

Appending models and their data requirements (ID model type a) + b) + c) + d). Detailed data which depends on the spatial dimension to the water system modelling (1D-2D-3D) and temporal scale (monthly, daily, hourly, sub-daily) series of hydrodynamic data, sediments, mercury species concentrations, environmental conditions, bioaccumulated matter in biotic species.

•

and surface soils of which 72.5 % correspond to gold miningrelated activities. The remaining percentage is attributed to activities or sectors such as ferrous-metal and non-ferrous-metal mining, cement production, and industrial waste among others [13]. In Colombia, gold production is spread among extensive regions of the country. However, the official data for production of gold by department for 2012, published by the Colombian Mining Information System (Sistema de Información Minero Colombiano - SIMCO), indicate that almost 85 % of the national production is focused in three departments: Antioquia accounts for 42 %; Choco is responsible for 37 %; and Bolivar holds 6 % [50]. Additionally, it is estimated that approximately 40 % of the national gold average annual production is concentrated in a region of the Antioquia department known as the Lower Antioquean Cauca, which groups the Segovia, Taraza, Caceres, El Bagre, Zaragoza, Caucasia and Nechí municipalities [51]. It is important to consider that in Colombia, Artisanal and Small Scale Gold Mining (ASGM) activities accounts for approximately 40 % of the total national gold exploitation. In general, ASGM is developed through practices that are inadequate in terms of production, and without regard to environmental impacts, social 88


responsibility or the occupational health of the mining workers [13]. The proportion of this type of gold mining activities, which is developed under no regulation or control in productive and environmental terms, can imply that the polluting loads released or poured into the environment derived from the ASGM can be even higher than those previously reported, which further magnifies the environmental mercury contamination problem in the country. In Colombia, the hydrological basins most impacted by direct mercury discharges associated with gold mining according to the 2014 National Study on Water [49] are: the Magdalena River – Morales arm branch (Bolivar department), the Nechí River and the main stream tributaries of its lower basin (Antioquia department – Lower Cauca region), the main stream tributaries of the Cauca River in its middle and lower basins such as the Taraza River and the Man River (Antioquia department – Lower Cauca region), the main stream tributaries of the middle and lower basins of the San Juan River (Choco department), and the San Jorge River in its lower basin (Cordoba and Sucre departments – Mojana region). Most of these basins, like the Lower Antioquean Cauca region, coincide with intensive gold production activities. In biophysical terms, the region referred to as Mojana region (Figure 2) is delimited by the river delta formed in the north by the Loba arm branch of the Magdalena River, in the west by the lower basin of the San Jorge River, and in the east by the Cauca River. This river delta forms in its interior and surroundings a floodplain that forms a large system of wetlands and consequently an intricate network of swamps, creeks, lagoons and marshes [52]. In political terms, this region falls under the jurisdiction of eleven municipalities pertaining to four departments; Nechí (Antioquia department), Magangué, Achí, San Jacinto del Cauca (Bolivar department), Ayapel, San Marcos (Cordoba department), 89

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Guaranda, Majagual, Sucre, Caimito, San Benito Abad (Sucre department). It is important to consider that the Mojana region shares its southeastern border with the Lower Cauca mining region, which is located in the lower-middle basin of the Cauca River, in hydrological terms.

Figure 2. The Mojana region area, hydrosystem and Lower Cauca gold mining production region municipalities.

In the fluvial morphology and hydrological context, the floodplains serve the function of accumulating material produced and transported in its river basins (solids, sediments, detritus, debris, and organic matter among others) [52]. Moreover, they serve to regulate floods in the river basins and promote the mitigation of their effects. Apart from these regulation ecosystem services, that are highly important in a river basin of great size like the Magdalena – Cauca river where is located the Mojana region, these floodplain macro-ecosystems are globally known for their ecological importance in terms of biodiversity and due to the variety of ecosystem services they provide to humanity and mainly to local communities [53]. 90


As a floodplain in the Magdalena-Cauca river basin, the Mojana region intrinsically provides the ecosystem service of retaining, processing and accumulating solids, nutrients and energy flows generated and transported through its multiple streams [54]. This natural condition is combined with its spatial proximity to the Cauca river upstream territories where mercury polluting loads produced in gold mining are discharged into the environment; i.e., the Lower Antioquean Cauca, the San Lucas mountain range (eastern natural boundary of the Mojana region - Bolivar department), and the Mojana region itself. A large portion of these polluting loads is transported through the lotic water bodies of the basin due to hydrodynamic and sediment transport processes. These water bodies include the Magdalena (Loba arm branch), Cauca, Nechí, Caribona, San Jorge Rivers and also the great wetlands system of the Mojana region, where these streams have a direct connectivity and exchange flows, matter, and temporary or seasonal energy with the Mojana hydrosystem. The mercury loads are transported to the lotic and lentic water bodies of the Mojana region and this becomes the final destination of the mercury species, which can then transform via biogeochemical processes and can bioaccumulate in the hydrobiological species and biomagnify in their food chains. Research on the quantification of mercury in various environmental compartments of the Mojana region and their environmental and sanitary effects has been conducted approximately since the 1990s [4]. The results from these studies reveal that Mojana regionhas been a mercury contamination problem. As a consequence of those findings, with the development of local, national and global conscience regarding the toxicity of mercury, and after comprehending the magnitude of this environmental and toxicological problem, multiple studies and investigations have been carried out with different objects and scopes, all with the purpose of assess the mercury contamination problem in the 91

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Mojana region. In general, the scope of these studies falls within the eco-toxicological, environmental and public health fields, and establish mercury and methylmercury concentrations in various environmental compartments such as water, sediments, air and soil combined with biotic species from multiple levels of the food chain such as algae, plankton, aquatic vegetation, ichthyic species, amphibians, terrestrial species, cultivated rice and human urine, hair and blood [55]. The most representative investigations and studies references recommended to deepen the details of the mercury contamination problem in the Mojana region in the compartments mentioned and in local communities are: Marrugo (2015), Una Mirada a impactos del mercurio en Colombia [55]; Olivero-Verbel and Johnson (2002), El lado gris de la minería de oro [4]; and Pinedo et al. (2015), Speciation and bioavailability of mercury impacted by gold mining in Colombia [56]. These publications mainly show results about the various mercury measurements developed in the last few years in the Mojana and neighboring regions in multiple environmental compartments and human communities. Furthermore, they analyze their results in the light of references that point to the environmental problem and the threat to local public health, and they characterize the health effects on the local population that is directly exposed to the contaminated environment. In general, the results of these studies are a clear evidence of the mercury contamination problem in the aquatic environments of the Mojana region in direct connection with the use of this metal in human activities developed in the Magdalena – Cauca river basin such as gold mining, and the direct or indirect discharge or release of polluting loads into the environment in the surrounding upstream territories of its basin, which contaminate the water bodies of the complex hydrosystem of the Mojana region due 92


to mercury transport and fate phenomena. These polluting mercury loads transported and accumulated in the Mojana region hydrosystem subsequently could be transformed into mercury bioavailable species that are biomagnified in the food chains and exposed to the humans, becoming a threat for the health of the local human population.

Data requirements and perspectives for the development of surface water mercury models of transport, fate and transformation in the Mojana region The application of a model that simulates mercury dynamics in aquatic systems in the hydrosystem of the Mojana region implies much complexity due to the intrinsic features of this great floodplain formed by multiple lotic and lentic surface water bodies added to the connectivity and exchange of flows, matter and permanent or temporary energy with large rivers (Cauca River, the Loba arm branch of the Magdalena River and San Jorge River), and human activity. Thus, its characterization and the modelling of its hydrodynamics and sediment transport processes would require great economic, technical and institutional resources in order to assess the transport and fate of mercury species in the hydrosystem. However, regarding the hydrodynamic model, at present there is a regional hydrodynamic model that was developed by the national institution Fondo Adaptación in 2014 in the context of the project Plan de acción de intervención integral para la reducción del riesgo de inundaciones en la region de la Mojana (Intervention action plan for reducing the risk of floods in the Mojana region) [52]. The model was designed with the purpose of simulating scenarios of flood hazards in the region. Among its main features, it allows the spatial simulation of Mojana region lotic water bodies in one 93

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dimensional space, and of lentic water bodies and flooding areas in 2-Dimensional space, with spatial resolution of cell size of approximately 400 meters and a sub-daily time scale. The model was developed with the SOBEK software, which is a numerical modelling software package that allows the simulation of various hydrodynamic, biophysical, morphological and ecological processes. It was developed in the 1980s by the Deltares research and technology institute in Delft, Netherlands, and its main advantages include its documentation, support, continuous development and updating process, which incorporates state of the art advances in related fields of hydrodynamics and biophysics [57]. The existence of a hydrodynamic model for the Mojana region hydrosystem with the features mentioned represents a great advantage for undertaking a mercury dynamics modelling project in this aquatic ecosystem. The model could be used to simulate the transport and fate processes of mercury species, integrating a sediment transport model for mercury fractions linked to suspended matter combined with a quality sub-model for simulating the dynamics of mercury species dissolved in the water column. However, it is important to consider the data requirements for the application of this model of transport and fate of mercury species cited in Table 1, ID model type b), summarized as follows: for sediment transport (grain-size distribution of suspended and bottom sediments, concentration of total suspended solid fractions, dissolved solids and average density of the particles), concentrations of mercury species in the water column, specific polluting loads, and suspended sediment and bottom sediment loads. It is important to take into account that the data requirements will be greater if the mercury dynamics modelling project for the Mojana region hydrosystem intends to simulate transformation, bioavailability and bioaccumulation processes (ID model type c), d) and e) according to Table 1). 94


Even though various studies and measurements on mercury species in different compartments of the Mojana region water bodies, on hydrobiological species and other environment compartments (air, soil, humans, terrestrial animals), which they have been developed in the past few years, their main goal is to prove the environmental condition of mercury contamination in the region combined with the toxicological effects that imply risks for human health. Consequently, these studies, which have not been specifically developed to provide input data for modelling or for the calibration of the mercury dynamics model, lack the fundamental requirements to be used in a modelling project of the previously mentioned characteristics; i.e., the periodicity of the measurements according to hydrological regimes; the spatial coverage of the measurements according to the spatial scale of simulation of the model; the hydrodynamic, sediment transport and mercury species characterization in the boundary conditions of the model; and the specific polluting loads among other aspects. These requirements restrict the use of the information available produced in existing studies and investigations about the topic for the implementation of a transport and fate model or in more complex terms a transformation and bioaccumulation model in the Mojana region [10][12][36]. These limitations in the available and required data for the development of mercury dynamics modelling projects in the Mojana region point to the need of conducting specific activities to gather this information for the purpose of modelling, as is the case with the formulation of a monitoring plan requiring, at least, the sampling spatial distribution, the periodicity of monitoring campaigns, and mercury species, hydrodynamic, sediment and environmental variables to be sampled and measured, among other aspects, as well as an estimation of the logistics and economic, human and technical resources required for the development of such monitoring plan [12]. 95

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Other additional considerations to the data requirements for the implementation of a model like the one described and the implications related to the formulation of a monitoring plan are the economic, technical, human, space and time, institutional and other resources fundamentally represented in the technical capacity to develop the model with highly qualified technicians, investment in specific software and hardware and its corresponding technical support, sustainability of the project in time in economic terms, the institutional commitment and responsibility on the development of these models that represent significant economic investments for institutions or organizations that develop this kind of studies, which is why it is of the utmost importance to conduct a cost-benefit analysis of the various resources required for the application of the models, to specify the type of model that meets the object of the project, and the available resources prior to undertaking a mercury species transport, fate and transformation model in water bodies. However, it is necessary to further detail the present magnitude of the mercury contamination problem in the Mojana region hydrosystem. This would justify the need for undertaking a project with a model of this type, as well as build from existing inputs such as the technical, human and economic resources invested in the development of the hydrodynamic model of the core of eleven municipalities of the Mojana region conducted by the national institution Fondo Adaptación [52]. This would be a crucial step in the implementation of mercury transport, fate and transformation models in this important hydrosystem.

References [1] UNEP- Pirrone, N.; Mason, R. Mercury fate and transport in the global atmosphere. Dordrecht, The Netherlands: Springer. DOI, 10, 978-0, 2010. [2] UNEP, G. M. A. Sources.Emissions, Releases and Environmental Transport, UNEP Chemicals Branch, Geneva, Switzerland, 2013.

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[3] Gaioli, M.; Amoedo, D.;& González, D. Impacto del mercurio sobre la salud humana y el ambiente. Archivos argentinos de pediatría, 2012, 110(3), 259-264. [4] Olivero, J.; Johnson, B. El lado gris de la minería del oro: La contaminación con mercurio en el norte de Colombia. Editorial Universitaria. Colombia, 2002. [5] Pirrone, N.; Mason, R. Mercury fate and transport in the global atmosphere. Dordrecht, The Netherlands: Springer. DOI, 10, 978-0, 2009. [6] Gilmour, C. C.; Henry, E. A.; Mitchell, R. Sulfate stimulation of mercury methylation in freshwater sediments. Environmental Science & Technology. 1992, 26(11), 2281-2287. [7] Benoit, J. M.; Gilmour, C. C.; Mason, R. P.; Heyes, A. Sulfide controls on mercury speciation and bioavailability to methylating bacteria in sediment pore waters. Environmental Science & Technology. 1999, 33(6), 951-957. [8] Monperrus, M.; Tessier, E.; Amouroux, D.; Leynaert, A.; Huonnic, P.; Donard, O. F. X. Mercury methylation, demethylation and reduction rates in coastal and marine surface waters of the Mediterranean Sea. Marine Chemistry. 2007, 107(1), 49-63. [9] Drott, A.; Lambertsson, L.; Björn, E.; Skyllberg, U. Do potential methylation rates reflect accumulated methyl mercury in contaminated sediments?. Environmental science & technology. 2007, 42(1), 153-158. [10] Massoudieh, A.; Zagar, D.; Green, P. G.; Cabrera-Toledo, C.; Bombardelli, F. A. Modeling mercury fate and transport in aquatic systems. Advances in Environmental Fluid Mechanics.World Scientific, London, New York, Singapore, 2010. [11] Wang, Q.; Kim, D.; Dionysiou, D. D.; Sorial, G. A.; Timberlake, D. Sources and remediation for mercury contamination in aquatic systems—a literature review. Environmental pollution, 2004, 131(2), 323-336. [12] DTMC y SRWP. Modeling Mercury Fate, Transport, and Uptake in the SRW. Delta Tributary Mercury Council and SRW Program-Mercury Models Report. Appendix 4, 2002. [13] Programa de las Naciones Unidas para el Medio Ambiente (PNUMA). Sinopsis nacional de la minería aurífera artesanal y de pequeña escala. Programa de las naciones unidas para el medio ambiente – PNUMA ministerio de ambiente

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y desarrollo sostenible– MADS. Informe final. Bogotá. Diciembre de 2012. 204 p. [14] Tong, Y.; Zhang, W.; Chen, C.; Chen, L.; Wang, W.; Hu, X.; Wang, Q. Fate modeling of mercury species and fluxes estimation in an urban river. Environmental Pollution. 2014, 184, 54-61. [15] Brigham, M. E.; Wentz, D. A.; Aiken, G. R.; Krabbenhoft, D. P. Mercury cycling in stream ecosystems. 1. Water column chemistry and transport. Environmental science & technol. 2009, 43(8), 2720-2725. [16] Canavan, R. W.; Slomp, C. P.; Jourabchi, P.; Van Cappellen, P.; Laverman, A. M.; Van den Berg, G. A. Organic matter mineralization in sediment of a coastal freshwater lake and response to salinization. Geochimica et Cosmochimica Acta. 2006, 70(11), 2836-2855. [17] Hudson, R. J.; Gherini, S. A.; Watras, C. J.; Porcella, D. B. Modeling the biogeochemical cycle of mercury in lakes: The mercury cycling model (MCM) and its application to the MTL study lakes. Mercury Pollution: Integration and Synthesis. 1994, 473-523. [18] Shrestha, P. L. An integrated model suite for sediment and pollutant transport in shallow lakes. Advances in Engineering Software, 1996, 27(3), 201-212. [19] Gunneriusson, L.; Sjöberg, S. Equilibrium speciation models for Hg, Cd, and Pb in the Gulf of Bothnia and its catchment area. Hydrology Research. 1991, 22(1), 67-80. [20] Zhang, J.; Wang, F., House, J. D.; Page, B. Thiols in wetland interstitial waters and their role in mercury and methylmercury speciation. Limnology and Oceanography. 2004, 49(6), 2276-2286. [21] Žagar, D.; Knap, A.; Warwick, J. J.; Rajar, R.; Horvat, M.; Četina, M. Modelling of mercury transport and transformation processes in the Idrijca and Soča river system. Science of the Total Environment. 2006, 368(1), 149-163. [22] Ethier, A. L. M.; Mackay, D.; Toose-Reid, L. K.; O’Driscoll, N. J.; Scheuhammer, A. M.; Lean, D. R. S. The development and application of a mass balance model for mercury (total, elemental and methyl) using data from a remote lake (Big Dam West, Nova Scotia, Canada) and the multi-species multiplier method. Applied Geochemistry. 2008, 23(3), 467-481. [23] Officer, C. B.; Lynch, D. R. Bioturbation, sedimentation and sediment-water

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exchanges. Estuarine, Coastal and Shelf Science. 1989, 28(1), 1-12. [24] Radovanovic, H.; Koelmans, A. A. Prediction of in situ trace metal distribution coefficients for suspended solids in natural waters. Environmental science & technology. 1998, 32(6), 753-759. [25] Turner, R. R.; Lindberg, S. E. Behavior and transport of mercury in riverreservoir system downstream of inactive chloralkali plant. Environmental Science & Technology. 1978, 12(8), 918-923. [26] Fontaine, T. D. A non-equilibrium approach to modeling toxic metal speciation in acid, aquatic systems. Ecological Modelling. 1984, 22(1), 85-100. [27] Mackay, D.; Diamond, M. Application of the QWASI (Quantitative Water Air Sediment Interaction) fugacity model to the dynamics of organic and inorganic chemicals in lakes. Chemosphere. 1989, 18(7-8), 1343-1365. [28] Hudson, R. J., Gherini, S. A., Watras, C. J., &Porcella, D. B.Modeling the biogeochemical cycle of mercury in lakes: The mercury cycling model (MCM) and its application to the MTL study lakes. Mercury Pollution: Integration and Synthesis. 1994, 473-523. [29] D. Leonard et al., Water Air Soil Pollut. 1995, 80,519-528. [30] Knightes, C. D.; Ambrose, R. B. Evaluating regional predictive capacity of a process‐based mercury exposure model, regional‐mercury cycling model, applied to 91 Vermont and New Hampshire lakes and ponds, USA. Environmental Toxicology and Chemistry, 2007, 26(4), 807-815. [31] Kotnik, J.; Horvat, M.; Jereb, V. Modelling of mercury geochemical cycle in Lake Velenje, Slovenia. Environmental Modelling & Software. 2002, 17(7), 593611. [32] Knightes, C. D.; Ambrose, R. B. Evaluating regional predictive capacity of a process‐based mercury exposure model, regional‐mercury cycling model, applied to 91 Vermont and New Hampshire lakes and ponds, USA. Environmental Toxicology and Chemistry. 2007, 26(4), 807-815. [33] Brown, S., Saito, L.; Knightes, C.; Gustin, M. Calibration and evaluation of a mercury model for a western stream and constructed wetland. Water, air, and soil pollution. 2007, 182(1-4), 275-290.

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[34] Canu, D. M.; Acquavita, A.; Knightes, C. D.; Mattassi, G.; Scroccaro, I.; Solidoro, C. Modeling the Mercury Cycle in the Marano-Grado Lagoon (Italy). Models of the Ecological Hierarchy: From Molecules to the Ecosphere. 2012, 25, 239. [35] Henry, E. A.; Dodge-Murphy, L. J.; Bigham, G. N.; Klein, S. M.; Gilmour, C. C. Total mercury and methylmercury mass balance in an alkaline, hypereutrophic urban lake (Onondaga Lake, NY). Water, Air, and Soil Pollution. 1995, 80(1-4), 509-517. [36] Ambrose, R. B.; Wool, T. A. WASP7 Stream transport-model theory and user’s guide, supplement to water quality analysis simulation program (WASP) user documentation. National Exposure Research Laboratory, Office of Research .and Development, US Environmental Protection Agency, Athens, Georgia. 2009. [37] Carroll, R. W. H.; Warwick, J. J.; Heim, K. J.; Bonzongo, J. C.; Miller, J. R.; Lyons, W. B. Simulation of mercury transport and fate in the Carson River, Nevada. Ecological Modelling. 2000,125(2), 255-278. [38] Sonke, J. E.; Heimbürger, L. E.; Dommergue, A. Mercury biogeochemistry: Paradigm shifts, outstanding issues and research needs. ComptesRendus Geoscience. 2013, 345(5), 213-224. [39] Monperrus, M.; Tessier, E.; Amouroux, D.; Leynaert, A.; Huonnic, P.; Donard, O. F. X. Mercury methylation, demethylation and reduction rates in coastal and marine surface waters of the Mediterranean Sea. Marine Chemistry, 2007, 107(1), 49-63. [40] Bouchet, S.; Amouroux, D.; Rodriguez-Gonzalez, P.; Tessier, E.; Monperrus, M.; Thouzeau, G.; Grall, J. MMHg production and export from intertidal sediments to the water column of a tidal lagoon (Arcachon Bay, France). Biogeochemistry. 2013,114(1-3), 341-358. [41] Bhavsar, S. P.; Diamond, M. L.; Gandhi, N.; Nilsen, J. (2004). Dynamic coupled metal transport‐speciation model: Application to assess a zinc‐contaminated lake. Environmental toxicology and chemistry. 2004, 23(10), 2410-2420. [42] Gandhi, N.; Bhavsar, S. P.; Diamond, M. L.; Kuwabara, J. S.; Marvin‐ DiPasquale, M.; &Krabbenhoft, D. P. Development of a mercury speciation, fate, and biotic uptake (BIOTRANSPEC) model: application to Lahontan Reservoir (Nevada, USA). Environmental Toxicology and Chemistry. 2007, 26(11), 22602273.

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[43] Harris, R. C.; Snodgrass, W. J. Bioenergetic simulations of mercury uptake and retention in walleye (stizostedionvitreum) and yellow perch (percaflavescens). Water Quality Research Journal of Canada, 1993, 28(1), 217-236. [44] Parkhurst, D. L.; Appelo, C. A. J. User’s guide to PHREEQC (Version 2): A computer program for speciation, batch-reaction, one-dimensional transport, and inverse geochemical calculations. 1999. [45] Sunderland, E. M.; Mason, R. P. Human impacts on open ocean mercury concentrations. Global Biogeochemical Cycles, 2007, 21(4). [46] A.E. Bale, J. Environ. Eng., ASCE 2000,126,153-163 [47] Harris, R. C.; Bodaly, R. D. Temperature, growth and dietary effects on fish mercury dynamics in two Ontario lakes. Biogeochemistry. 1998, 40(2-3), 175187. [48] Thibaud, Y. Use of the Thomann model for interpretation of mercury concentrations in Atlantic fishes [methylmercury]. Aquatic Living Resources (France) 1992. [49] IDEAM. Estudio Nacional del Agua 2014. Bogotá, D. C., ISBN: 978-9588067-70-4, 2105. 496 pag. [50] Home page. Sistema de Información Minero Colombiano. http://www. simco.gov.co/ (accesedMarch 20, 2015). Unidad de Planeacion Minero Energetica (UPME) – Bogota D.C. 2015. [51] Fedesarrollo. Estudio sobre los impactos socio -económicos del sector minero en Colombia: encadenamientos sectoriales. Estudio preparado para la asociación de la minería a gran escala. Bogota, D.C., 2013. 69 pag. [52] Fondo Adaptación. Plan de acción de intervención integral para la reducción del riesgo de inundaciones en la región de La Mojana. Documento resumen. Bogota, D. C. 2015. 86 pag. [53] Vilardy, S.; González, J.A. (Eds.). Repensando la Ciénaga: Nuevas miradas y estrategias para la sostenibilidad en la Ciénaga Grande de Santa Marta. Universidad del Magdalena y Universidad Autónoma de Madrid. Santa Marta, Colombia. 2011,228 pag. [54] Fondo Adaptación - García, L. Funcionamiento del sistema natural y

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características de los ecosistemas del núcleo de once municipios de la región de La Mojana. Documento producto del contrato 003 de 2014 Fondo Adaptación. 2015, 102 pp. Publicación digital. [55] Marrugo, J. Una mirada a impactos del mercurio en Colombia. Capítulo del libro: El problema de contaminación por mercurio, nanotecnología: retos y posibilidades para medición y remediación. Ed: E. González, J. Marrugo, V. Martínez. Red Colombiana de Nanociencia y Nanotecnología. 2015. 205 pag. [56] Pinedo-Hernández, J., Marrugo-Negrete, J., &Díez, S. Speciation and bioavailability of mercury in sediments impacted by gold mining in Colombia. Chemosphere, 2015, 119, 1289-1295. [57] Deltares. D -WAQUser Manual.Simulation of multi-dimensional hydrodynamic and transport phenomena, including sediments.Deltares Institute .Delft – Holland. 2011, 478 pag.

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Remediation

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Nano modified clays, bioclays and bio-leaching for water and sediments remediation Natalia Porzionato, Luz M. Guz, Melisa Olivelli Gustavo A. Curutchet, Roberto J. Candal

W

ater and sediments contamination is one of the more dramatic environmental problems that humanity is facing now. The situation is worst in the undeveloped countries, with overcrowd cities and deficient water management. Heavy metals, pathogens and recalcitrant organic compounds are between the most dangerous pollutants found in those environments. Versatile and relatively cheap processes for water, waste water and sediments purification should be developed in an attempt to remediate the situation. In this chapter, a few examples of water and sediment contamination in Argentina are presented, as well as remediation alternatives based in the use of nano and bioclays.

Instituto de Investigación e Ingeniería Ambiental, CONICET, Universidad Nacional de San Martín, Campus Miguelete, 25 de mayo y Francia, 1650 San Martín, Provincia de Buenos Aires, Argentina. e-mail: [email protected] 105

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Introduction Water contamination is one of the main concerns in the entire world but in particular in undeveloped countries, were environmental control and regulations are lighter than in developed countries, or it application is more difficult due to social or economic constrains [1]. Surface water contamination is mainly consequence of the release of untreated Municipal or industrial effluents to water courses. Municipal effluents, in the cases where industrial effluents are separated from sewage waters, can be relatively easy treated by conventional biological methods. The main problem with sewage waters, beside the enormous volume of water to be treated in big cities, is the presence of the so called emergent contaminants. This group of contaminants include medicines, hormones, and other substances eliminated by human been that cannot be degraded in conventional biological plants. Industrial waste water represents a more complicated problem because some contaminants can be recalcitrant or even toxic to microorganisms. Consequently, conventional bio-treatment may be not enough to eliminate these pollutants. In these cases, different physicochemical methods are typically used which include coagulation/precipitation; oxidation, neutralization, etc. Physicochemical methods can be used alone or coupled with bio treatment and/or adsorption. In the case of coagulation/flocculation and adsorption, the production of solid waste containing concentrated amounts of contaminants can be a problem difficult to solve. Subsurface water can be contaminated by sewage water, industrial effluents, chemicals and oil spilling. But also there are natural pollutants as arsenic, which may be presented in high concentration in subsurface water. In situ remediation of contaminated subsurface water is a difficult task and is one of the more important and interesting modern challenge for environmental and hydraulic engineers. 106

Nano modified clays, bioclays and bio-leaching

When conventional water treatment fails, nanotechnology is one of the modern tools that researchers and engineers can use to resolve complicated problems. Nanomaterials display newest and powerful capacities as adsorbents or catalysts, react faster than regular size materials, can penetrate or migrate to places unavailable for other systems, etc. Newly developed nanomaterials and nano-devices are available or are being studied to resolve pollution problems associated with recalcitrant compounds, metals in water, emergent contaminants, disinfection and even more. In this chapter some typical examples of different heavily polluted systems in Argentina will be presented as examples of contamination in undeveloped countries, followed by the discussion of different treatment process that involve the use of clays modified in the nanoscale, and bioremediation of metal contaminated sites using native bacteria consortia.

Water contamination in Argentina: a few typical examples Metal contamination in water and sediments at Reconquista River basin: the case of the José Leon Suarez Channel The Reconquista is one of the more polluted rivers in the country and is emblematic of environmental problems. This river receives contributions from storm sewers and streams that run across zones of high population density and serious environmental problems (hyper-degraded territories). Many of the tributaries of this basin receive domiciliary and industrial contaminating loads [2]. One of them is the José León Suarez channel (JLSC). This stream runs tubing as a rainwater collector and leaves open at “La Cárcova” 107

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neighborhood situated approximately 4.5 Km. upstream of its drainage at the Reconquista river, as can be seen in Figure 1.

Figure 1. Reconquista River, José León Suarez channel and adjacent neighborhoods. © 2016 Google.

This neighborhood has characteristic of extreme poverty, lack of drinking water services, sewer and energy, and its population is highly exposed to contaminants carried by the channel and those accumulated in the sediments. Figure 2 shows images of drainages discharging in the channel, very close to the houses. Although the JLSC is originated by the confluence of various storm sewers, it pulls high pollution levels as a consequence of irregular loads. However, due to the high water self-purification capacity of these streams, acceptable level of organic contamination (oxygen chemical demand less than 50 mg L-1) reaches the Reconquista River, located only 5 km downstream. The self-purification rate overcomes the normal degradation rate of


a)

b)

Figure 2. a) example of drainage that empties into the cannel; b) view of the channel at the border of the neighborhood.

organic matter due to the processes of sedimentation. This process leads to the incorporation of organic matter and other pollutants (as metals) into sediments. The high concentration of organics in the sediments generates a high benthic oxygen demand, setting an anoxic environmental condition suitable for the biocatalyzed formation of sulphides. Under these low redox potential conditions, most of heavy metals might precipitate as poorly soluble sulphides and hydroxides or are adsorbed to the different mineral components of the sediment matrix. The concentration of metals in the sediments corresponding to different points along the JLSC indicated in Figure 3 is shown in Table 1. Table 1 also displays the concentration of the same metals in the sediments of Roggero Dam (a dam placed at the nascent of Reconquista River), Hidalgo channel (H) (a channel close to JLSC), and in the Reconquista River (R) in an area close to where JLSC and H discharge it waters at Reconquista River. Data presented in 109

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Figure 3. samples sites along the JLSC. Imágenes © 2016 Google Datos del mapa © 2016 Google.

Table 1. Metal concentration along the JLSC (see Figure 3). Total content of Cu, Cr, Zn. Comparison with Hidalgo (H), Reconquista (R) and Roggero Dam (C); concentrations in [mg Kg-1]; (ND): no determined.

Site

Metal Copper

Chromium

Zinc

2

247

8

181

6

350

118

1570

8

220

14

650

9

15

7

330

11

ND

ND

800

12

120

56

340

H

115

105

450

R

630

2540

1020

C

6

Level 3 Metal/Metalloid

Level 1

Level 2

Level 3

As

3

9

33

Cd

0.6

0.9

10

Zn

100

271

540

Cu

28

50

110

Cr

26

55

110

Ni

16

35

75

Pb

23

42

250

These heavily contaminated sediments are very dangerous because if they are expose to atmospheric oxygen, biocatalized processes can produce an incontrollable leaching of dissolved heavy metal to the water course. This phenomenon happens when the sediments are removed by dragging and piled at the side of the river. Water 111

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Water contamination by textile dyes The presence of textile dyes in water is beside an example of landscape contamination, a serious risk for the biota and human being [3-5]. The use of this type of compounds notably increased during the last decades in the entire world [1]. However, due to legislation and strict controls on industrial effluents implemented in Developed Countries, the economic crisis and the need to diminish production costs, a significant fraction of that intensively use textile dyes was relocated to Undeveloped Countries. Figure 4 shows the evolution of dyes consumption during the last decades in different areas of the world.

Figure 4. Evolution of dye consumption during the last decades. Adapted from Hessel et al 2007 [1].

Textile industries in Argentina are concentrated in different areas of the country. In particular, San Martín County has an important concentration of textile factories. The effluents of these industries are discharged in the waters of Reconquista River and other streams with different degrees of treatment [2, 6]. Figure 5 shows an image of the waters of Medrano stream colored in blue due to the presence of textile dyes. 112


Figure 5. Medrano stream colored waters (San Martín, Buenos Aires province).

This contaminated waters discharge in the Rio de la Plata, producing important damages in coastal areas close to the river, increases the cost of water purification, affect the biota and hinder the recreational use of the river. Most of the textile companies are medium or small size and need economical alternatives for waste water treatment.

Nano-modified montmorillonite for water decontamination Clays, and specially laminated clays as bentonites, are useful materials for water treatment due to its high quality as sorbents and supports for catalysts. Besides, clays are usually cheap and abundant materials widely distributed all around the world. 113

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Clays are composed by tetrahedral layers of silica and octahedral layers of gibbsite. Figure 6a shows schemes of both structures. The central Si(IV) atoms of the tetrahedral can be replaced by Al(III) or Fe(III) atoms, in a phenomenon named isomorphic substitution. As a consequence, the structure acquires negative charge. In the case of the gibbsite layer the same phenomenon occurs with Al(III), which can be isomorphically substituted by Mg(II) or Fe(II). This layer also has a delocalized negative charge. The tetrahedral and octahedral layers join one which other through the apical oxygen of the tetrahedral layer with the OH of the octahedral layer. Depending on the quantity and distribution of the different layers clays can be classified in different families as 1:1; 2:1 and 2:1:1. The unit formed by one or two tetrahedral layers with one octahedral layer are called sheets and the clays are called bilaminar (1:1, T:O) or three-laminar (2:1, T:O:T). Between each sheet there is an inter-laminar space with intercalated alkaline (Na, K) or alkaline-earth (Ca, Mg) cations, that compensate the intrinsic negative charge of the sheets. Figure 6b shows the scheme of a three-laminar clay, with the interlaminar space and the cations located in this space. These clays are typically called smectites as, for example, montmorillonite (MMT) [7, 8]. Several laminar clays, as montmorillonite, can exchange the interlaminar cations by others inorganic or organic cations. Thanks to these characteristics, MMT is a versatile material that can be used to produce other simple or sophisticated new materials as sorbents, catalyst, polymers modified, etc. The exchange of the interlaminar cations by others with catalytic activity can produce a catalyst [9-14]. The exchange by organic cations (as quaternary ammoniums) can modify the hydrophilicity of MMT, leading to the production of sorbent for organics (as oil, solvents, etc)[15]. The sorption and catalytic properties of modified MMT can 114


Figure 6. a) Scheme showing the structure of the octahedral (Gibbsite) and tetrahedral (Silica) and its arrangement in layers.

Figure 6. b) Scheme of a three-laminar clay (2:1) with the interlaminar space. (Modified from [8]). 115

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be combined to obtain materials that can work as sorbents and catalysts at the same time or sequentially. This type of materials is very promising for environmental remediation. One example is the use of MMT modified with iron oxide nanoparticles for sorption of a cationic dye, followed by oxidation by photo-Fenton process [16]. Fe(III) can be incorporated as iron oxo-hydroxide nanoparticles (FeOx-NP) by the treatment of a suspension of MMT in acetone with FeCl3, followed by lyophilizing. Figure 7 shows images taken by scanning electron microscopy of pure MMT and Fe(III) modified montmorillonite (MMT-Fe). The images clearly show thin flakes of MMT and the FeOx-NP deposited on the MMT-Fe flakes [17].

Figure 7. SEM images of montmorillonite (MMT) and montmorillonite modified with Fe(III) oxo-hydroxide nanoparticles (MMT-Fe). Reprinted from Guz et al. 2014 [17], with permission from Elsevier.

Fe(III) was also located in the interlayer space as determined by X-Ray diffraction analysis (XRD). Figure 8a shows the XRD pattern of MMT and MMT-Fe. The diffraction peak centered at 2θ = 6.51° correspond to and interspace of 13.6 A between sheets in the MMT, while in the case of MMT-Fe the peak shifted to 5.5 A corresponding to an interspace of 16.5 A. These results indicate that FeOx-NP were also located between the sheets increasing the 116


interspace. The presence of FeOx-NP at the surface of the MMT-Flakes also modified the electrophoretic behavior and the surface charge of montmorillonite. As shown in Figure 8b, the Z potential (estimated by measurements of electrophoretic mobility) became notably less negative as pH decreased in the case of MMT-Fe. This phenomenon was consequence of the higher isoelectric point of iron oxide (close to pH 8.0), that modified the overall behavior of MMT-Fe with respect to MMT. These result indicated that the surface charge of MMT-Fe was notably more positive than MMT at pH lower than 6.

Figure 8. a) XRD pattern of MMT and MMT-Fe; b) Z potential vs pH for MMT and MMT-Fe. Reprinted from Guz et al. 2014 [17], with permission from Elsevier.

Adsorption of crystal violet (CV) on MMT was different than on MMT-Fe, as shown in Figure 9. The adsorption isotherms of CV on MMT or MMT-Fe were fitted in terms of a two sites Langmuir model. As can be seen in Table 3, the maximum amount of CV adsorbed by MMT (Qmax1 + Qmax2) was approximately 500 mgdye/ gclay, notably higher than for MMT-Fe: 200 mgdye/gclay. These results were consequence of the change in surface charge and the presence of FeOx-NP in the intersapace. 117

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The analysis of samples of CV adsorbed at MMT and MMT-Fe by Small Angle X-Ray Scattering (SAXS), allowed the precise Table 3. adsorption constants obtained after fitting the data with two sites Langmuir Model. Qmax: maximum amount of CV adsorbed on MMT or MMT-Fe (mgdye/gclay). Kd: desorption constant (Guz et al 2014, [17]).

R2

Qmax1

Kd1

Qmax2

Kd2

MMT

0.9519

327± 28

0.01± 0.005

174±26

12±6

MMT-Fe

0.9321

104±13

0.08±0.02

87±13

1.4±0.2

Figure 9. Adsorption isotherms of CV on MMT and MMT-Fe. The lines correspond with the fitting using one site or two sites Langmuir models. Reprinted from Guz et al. 2014 [17], with permission from Elsevier.

determination of the interlaminar distance. As shown in Table 4, the presence of CV notably increased the interlaminar space in bothe MMT and MMT-Fe. In the case of MMT the distance increased from 0.29 to 0.49 nm for a CV initial concentration in water in the range 0–150 ppm; however, for concentration higher than 250 ppm the interlaminar space increased to 0.98 nm. This behavior can 118


be consequence of dimer formation by CV. The dimeric moieties entered into the interlaminar space increasing the distance even more than in the case of monomeric CV. Similar phenomenon was observed in the case of MMT-Fe. However, in this case the interlaminar distance decreases at low CV concentration due to the exchange with the FeOx species located in the interlaminar space. This result indicated that the Fe(III) species were mobile and could be involved in ion exchange processes. Recalcitrant organic compounds can be totally or partially oxidized by photo Fenton process. Equations 1-5 illustrates the main reactions that take place during Fenton (1-3) and photoFenton process (4-5). Table 4. interlaminar space determined by SAXS.

[CV]1 MMT MMT-Fe Fe(II) + H2O2 → Fe(III) + H2O2 → Fe(III)-OOH2+ → Fe(III) + H2O → Fe(OH)2+ + hν →

0 ppm 0.29 nm 0.58 nm

250 ppm 0.98 nm 0.96 nm (1) (2) (3) (4) (5)

Hydrogen peroxide (H2O2) is the the typical oxidant used in Fenton and photo Fenton process. Fe(II) catalyze H2O2 decomposition in HO* radicals, which are powerful and no specific oxidants. In the dark the reduction of Fe(III) by H2O2 to regenerate Fe(II) is quite slow and controls the rate of the process. In the presence of light (λ < 400 nm) the product of Fe(III) hydrolysis (reaction (4)) generates Fe(II) and the overall reaction speed up. The optimal pH 119

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for these reactions is 3.0. Lower pHs decreases reaction rate and at higher pH Fe(III) precipitates. In the case of MMT-Fe, the FeOx-NP can act as catalyst for Fenton and photo-Fenton process. Figure 10 shows the degradation of CV adsorbed on MMT or MMT-Fe exposed to photo-Fenton treatment CV was completely adsorbed on MMT or MMT-Fe before the oxidation treatment was applied. In the case of MMT without added iron, some activity was observed, due to the catalytic action of the iron naturally present in the clay. After incorporation of Fe(III) or Fe(II) to the system with CV adsorbed on MMT, degradation rate was notably improved. However, the best result was obtained with MMT-Fe. These results indicate that Fe(III) present as FeOx-NP at the surface and inside MMT-Fe acted as an effective catalyst for photo-Fenton process. Figure 11 shows the evolution of total organic carbon in solution

Figure 10. Temporal evolution of CV adsorbed on MMT or MMTFe during photo-Fenton treatment. [H2O2] = 50 mM, pH 3, [Fe(II)] y [Fe(III)] = 0.5 mM, MMT y Fe-MMT 3 g/L, [CV] = 0.120 mM, T = 25°C, luz visible: 100 W/m2. Guz L., et al., 2014, [17] with permission from Elsevier. 120


(TOC) during photo-Fenton process. At the beginning of the treatment TOC increased due to the release of organic matter to the solution. This phenomenon is consequence of the partial oxidation of the CV adsorbed at the clay, followed by the release of the byproducts to the solution. The organics were further mineralized as the treatment advanced. The results presented above clearly show the feasibility to use

Figure 11. TOC evolution during photo-Fenton treatment. Similar conditions than in Figure 10. Guz L., et al., 2014; [17] with permission from Elsevier.

advanced oxidation, in particular photo-Fenton process, for the destruction of contaminants adsorbed at clays. In this way, the contaminants are not only removed from solution but completely eliminated, avoiding the complex problem of final disposition of dangerous pollutants. 121

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Bio-modified montmorillonite for water decontamination MMT is a good adsorbent for positive metal cations thanks to its intrinsic negative charge [18]. Metals cations can replace the cations naturally located in the interlaminar space (typically Na+, Ca2+, etc), but can also be strongly adsorbed at the surface [1921]. The equilibration is typically reached after few minutes of contact, being this process potentially useful for the treatment of waste waters containing metals (effluents from metallurgy, galvanoplastic companies, etc) [21]. Microbial biomass is also a very useful adsorbent for metals in water. Biomass is intrinsically negative due to the composition of the cellular membranes, which is rich in carboxylic functional groups. Researchers all around the world had demonstrate de feasibility of microbial biomass adsorbents in wastewater treatment [22]. In particular fungi are very interesting because they can growth fast on different surfaces at low cost [23, 24]. This two families of adsorbents have the advantages of being low cost materials with a relatively high adsorption capacity for metals. However, to increase its efficiency, they are used suspended in the wastewater to be treated. The principal drawback is the separation from the liquid phase, this process can be time consuming and expensive, particularly for clays [25]. It was recently demonstrated that the combination of clays with microbial biomass produces biopolymers with better coagulation properties and improved adsorption abilities for metals in water [23]. This is an example of the composition of two materials with synergically improved properties. Uranium, in the form of highly soluble U(VI), is a dangerous contaminant present in wastewaters produced during preparation 122


of nuclear fuels [26, 27]. It was recently demonstrated the feasibility to use adsorption on biopolymers prepared by combination of MMT with fungi as a way to remove U(VI) from water [23]. Biopolymers (BMMT) were prepared by growing biomass in the presence of MMT. Two uranium resistant fungi genus were used: Aphanocladium sp. (Apha sp.) and Acremonium sp. (Acre sp.). Figure 12 shows SEM images of Acre sp. alone, where the typical fibrous structure can be observed, and MMT with 1% of biomass (BMMT), where it can be noted that fungi growth intimately bound to the MMT flakes. Table 5 shows the biomass content of the different BMMT, and the H+ consumption capacity of MMT, Acre sp., Apha sp., and the different BMMT. The increment of the proton consumption capacity of BMMTs, with respect to that of MMT and biomass alone, demonstrates that clay biopolymers have an advantageous structure for chemisorption processes.

Figure 12. SEM images of Acre sp. (left) and BMMT (Acre sp.) composite (right).

Figure 13 shows adsorption isotherms of U(VI) on different adsorbents. As can be seen in the figure, the adsorption curves correlate with a sigmoidal model. This behavior is consequence of the heterogeneity of sites where the adsorbate can bind on the BMMT. Clays have interlayer spaces and external sites, and 123

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Table 5. MMT and BMMT biomass content and H+ exchange capacity (Olivelli et. al., 2013, [23]).

Sample

Biomass content (mg/g MMT)

mmol H+/g

(mmol H+BMMT– mmol H+MMT)/g biomass

MMT

0.6

Acre sp.

1.2

1.2

Apha sp.

1.2

1.2

BMMT 1% Acre sp.

237

1.2

2.5

BMMT 5% Acre sp.

57

0.9

5.2

BMMT 1% Apha sp.

257

1.3

2.7

BMMT 5% Apha sp.

73

0.95

4.7

biomass surface functional groups could provide a larger quantity of metal binding sites and greater affinity to the system [23]. The results presented in Figure 13 clearly show that the composite biomaterials displayed improved adsorption capacity for U(VI).

Figure 13. UO22+ adsorbed on BMMT and respective controls. A: BMMT=MMT+Apha; B: MMT+Acre sp. BMMT in 1% w/v. Reprinted from Olivelli et. al., 2013 [23], with permission from Elsevier 124


Bio-remediation of Sediments The sediments of the rivers heavily polluted with organic matter are expose to anoxic conditions, as was exemplified above with the cases of Reconquista River and JLS channel. Under these low redox potential conditions, most of heavy metals might precipitate as poorly soluble sulfides and hydroxides or are adsorbed to the different mineral components of the sediment matrix. The anoxic environment maintains metals in a low bioavailable state as long as low redox potential conditions are not altered. However, reduced compounds of sediments such as sulfides tend to oxidize due to redox processes biocatalyzed by sulfur oxidizing bacteria (SOB) when changes in the redox conditions occur [28-30]. This could be caused by exposition to oxygen, either by dredging operations or by drying effects (condition caused by a decrease in the water level). This phenomenon could impact directly on the environment because of the sediment acidification and the heavy metal release into the water column while their bioavailability seems to be incremented [31-36]. These processes have been intensively studied through static assays [37] and dynamic experiments of anaerobic sediments oxidation systems by re-suspension or desiccation [29, 33, 34, 36, 38-40]. The removal of heavy metals from contaminated sediment needs to be a priority as regards safer sediment managing because the risk of metal releasing into ground or surface water (and consequently the incorporation of heavy metal into the food chain). The oxidation/acidification processes described above have, under controlled conditions, the potentiality of recovering valuable metals from polluted sediments [40, 41, 42]. This set of processes is called bioleaching and was studied as a tool for the remediation of heavy polluted sediments in both agitated batch and in bioheaps systems. The reactions involved in the bioleaching processes could be developed by different mechanisms: acidic 125

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(interchange reactions or dissolution of acid soluble phases such as carbonates), oxidizing (metals bounded to organic matter or sulphides) or reducing (metals bounded to Fe(III) and Mn(IV) oxides). Most of these mechanisms are based on the biocatalyzed oxidation of the different phases of the sulphur content of the sediment. In these processes, the products (Fe(III), sulfuric acid and polithionates) of acidophilic bacteria metabolism (Acidithio bacillus ferrooxidans and Acidi thiobacillus. thiooxidans) act as reaction agents [28, 43]. The specific mechanisms of sulfide and sulfur oxidation by these species have been extensively studied [4447]. Beolchini et al. (2015) [48] reviewed the main contributions to the study of bioleaching of sediments and concluded that sediment bioremediation technologies have very site-specific effects. The particular characteristics of each sediment are strongly determinants of the kinetics and efficiency of metal extraction by bioleaching [49-52]. Bioleaching assays were performed on sediments from JLS channel (see Table 1) using stirred batch and bioheaps systems. Experiments in stirred batch were realized using 5, 10 and 15 % (w/v) pulp density (PD) suspended in 0K medium, inoculated with Acidithiobacillus thiooxidans and Acidithiobacillus thiooxidans DSM11478 in a concentration of 1.2×107 cell ml-1, with or without 5% added sulfur. Figure 14 a, shows the percentage of zinc extracted from the sediments and Figure 14 b the evolution of pH. Clearly, PD and sulfur addition have a key rol in zinc extraction and pH evolution, been pH and %Zn extracted intimately related. All systems without added sulfur showed a drop of 1.5 pH unit and an increment of the oxidant-reduction potential (ORP) close to 300 mV in the same time period. Sulfate concentration in these systems (data not shown) rose close to 800 mg L-1 in all conditions, suggesting that acidification is partially due to sulfides oxidation. In these cases, samples with 5% PD showed a final extraction of 40% of total Zn while those with 10 and 15% PD 126


showed an extraction of only 20% of total Zn. After the 28th day there was a notable increment in the percentage of extracted zinc. This phenomenon was associated with the drop of pH to values close to 5, that helped to solubilize zinc, being responsible for the increment in zinc extraction.

Figure 14. Bioleaching assay in batch mode with 5, 10 and 15% suspended sediment in 0K medium. (a): Percentage of Zn extracted; (b) reached pH values in systems with and without sulfur added. PD: pulp density, S: sulfur. Reprinted from [53] with permission from ACS.

In systems with added sulfur, a clear relationship between the extracted zinc and pH drop was observed. In systems with the lower sediment pulp density (5%), acidification produced by sulfur biocatalyzed oxidation was faster due to the poorer amount of neutralizing compounds provided by the sediment and because diffusional limitation was not important. These systems came close to a final extraction of 60-80% of initial zinc and reached pH values between 2 and 4. Systems with 10 and 15% PD showed an increment of ORP rise rate, zinc extraction rate and a pH drop at the 28th day. These results suggest the potentiality of the bioleaching process to extract metals from contaminated sediments amended with elemental sulfur in order to ensure a drop of pH value enough to 127

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allow metals stabilize in solution. The pulp density and the ratio of added sulfur determine acidification kinetics and the final pH. Given the huge volumes of sediment to be treated, performing a leaching process in stirred batch system could be economically not feasible. An alternative is the application of, bioheap leaching systems. The bioleaching heap system was compiled into Polyethylene terephthalate (PET) reactors of 12 cm high and 6 cm diameter. A first conditioning step was achieved to improve drainage and aeration. Conditioning was achieved using 5% w/w perlite addition to the sludge. Then elemental sulfur was added (1%, 2% 5% w/w) in the mixture in order to increment the electron source for sulfur oxidizing bacteria. One of the systems was just dusted superficially with 5%w/w sulfur. 100 g (dry weight) of conditioned sediment was used in each system, the bulk density was 280 g l-1. All the systems were inoculated with a mixture of Acidithiobacillus ferrooxidans (DSM 11477) and Acidithiobacillus thiooxidans (DSM 11478) previously suspended in 0K medium. A two stages irrigation methodology was performed; the aim of the first stage or acidification step was promoting the acidification generated by the oxidation of sulfides and elemental sulfur, while in the second step (or washing step) the aim was to drag all the soluble metal that have been released into the microenvironment. In the acidification step all the systems were irrigated with 50 ml of distilled water each 72-h. Before being recycled as irrigation water, leached water was collected and completed with distilled water until 50 ml. Aliquots of irrigation water were taken periodically to monitor pH, Zn and Cu content. Next, in the washing step, systems were washed 500 ml of distilled water. Figure 15 shows the percentages of copper and zinc extraction, and the final pH reached in different bioheaps. The final pH attained in systems without or with 1% added sulfur remained close to neutral, 128


Figure 15. Total percentages of copper and zinc of the bioheaps leached out, and final pH of the solutions for each system studied: No sulfur added, 1%, 2%, 5% of sulfur and 5% of superficial sulfur. S: sulphur.

whereas in systems with 2%, 5% sulfur and 5% surface sulfur the final pHs were 4.4, 2.9 and 2.3, respectively. The percentage values of zinc and copper extracted showed great differences between each system and can be associated with the final pH. In systems without and with 1% added sulfur, no net leaching was observed. This phenomenon could be attributed to the pH conditions which were close to neutrality during all the assay. In the others systems a significant leaching of zinc was observed, reaching 71% of efficiency in the case of the system with superficial scattered sulfur and between 36-53% for systems with 5% mixed sulfur. Meanwhile, only for both systems with 5%, a significant leaching of copper was detected. In general, system with 5% surface scattered sulfur showed higher extraction than system with 5% mixed sulfur. The best oxygen availability to perform sulfur oxidation in this system seems to be the main cause of these differences. 129

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Conclusions The fate of pollutants in water strongly depends on the physicochemical and microbiological characteristics of the water course. In the studied case, heavy metals concentrate in the sediments while organics are partially degraded by microbiological activity. The anoxic condition of the sediments leads to the production of stable sulfides, carbonates or reduced metal oxides. The concentration of metals in the more polluted areas is over the acceptable limits. Combination of clays with oxide nanoparticles or biomass produce suitable adsorbents for dyes, metals, etc. The combination of the catalytic properties of iron oxide nanoparticles with the adsorption capacity of clays, produced a material that can be used to separate contaminants from water and to catalytically oxidize the adsorbed pollutant (by photo-Fenton process). Bio-clays display superior adsorption capacity than the separated precursor materials, being useful for the separation of dangerous soluble metal species as U(VI). Bioleaching processes assisted by acid resistant bacteria is an alternative for the remediation of sediments heavily polluted with metals. This approach allows a controlled release of the metals that, potentially, can be recovered by appropriate chemical treatment. Environmental pollution is a drama that affect all the humanity but hit harder on undeveloped countries. In this chapter we presented different approaches that can be used to mitigate the problem. But the awareness of the people, leaders and authorities, about the environmental and social risk that is facing the Earth is strictly necessary for the well-being of the future generations.

Acknowledgments The Authors gratefully recognize the support given by Universidad 130


Nacional de San Martín, Ministerio de Ciencia y Tecnología, Consejo Nacional de Investigaciones Científicas y Técnicas, and Agencia Nacional de Promoción de Ciencia y Tecnología. RJC and GAC are researchers of CONICET.

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important interplay. Hydrometallurgy. 2006, 83, 153-166. [44] Viera, M.; Curutchet, G.; Donati, E. A combined bacterial process for the reduction and inmobilization of chromium. Int Biodeter Biodegr. 2003, 52, 31-34. [45] Rohweder, T.; Sand, W. Mechanisms and biochemical fundamentals of bacterial metal sulfide oxidation in: Donati ER and Sand W (Eds) Microbial Processing of Metal sulfides 2007. 53-58. [46] Vera, M.; Schippers, A.; Sand, W. Progress in bioleaching: fundamentals and mechanisms of bacterial metal sulfide oxidation-part A. Appl. Microbiol. Biotechnol. 2013, 97, 7529-7541. [47] Watling, H.R., Bioleaching of a low grade copper ore, linking leach chemistry and microbiology. Minerals Engineering, 2014. 56, 35.44. [48] Beolchini, F.; Rocchetti, L.; Fonti, V.; Dell’ Anno, A. Consequences of anaerobic biotreatments of contaminated sediments on metal mobility. International Journal of Environmental Science and Technology. 2015, 12 (7), 2143-2152. [49] Calmano, W.; Hong, J.; Förstner, U. Binding and mobilisation of heavy metals in contaminated sediments affected by pH and redox potential. Water SciTechnol. 1993, 28, 223–35. [50] Löser, C.; Zehnsdorf, A.; Görsch, K.; Seidel, H. Remediation of heavy metal polluted sediment in the solid bed: Comparison of abiotic and microbial leaching. Chemosphere. 2006, 65, 9-16. [51] Löser, C.; Zehnsdorf, A.; Hoffmann, P.; Seidel, H. Remediation of heavy metal polluted sediment by suspension and solid-bed leaching: Estimate of metal removal efficiency. Chemosphere. 2007, 66, 1699-1705. [52] Zehnsdorf, A.; Seidel, H.; Hoffmann, P.; Schlenker, U.; Müller R.A. Conditioning of sediment polluted with heavy metals using plants as a preliminary stage of the bioremediation process: a large-scale study. J Soils Sediment, 2013, 13, 1106-1112 [53] Porzionato, N.; Tufo, A.; Candal, R.; Curutchet, G. 2016 Metal Bioleaching from anaerobic sediments from Reconquista river basin (Argentina) as a potential remediation strategy. Environm Sci Pollut Res. DOI 10.1007/s11356-016-6717-y (in press).

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Biotechnological synthesis of silver nanoparticles using phytopathogenic fungi from cocoa Raquel Villamizar

T

he objective of this research was to explore the ability of native fungi isolated from cocoa crops, to biosynthesize nanoparticles. Once standardized, the nanoparticles were characterized such as UV-Vis spectrophotometry and scanning electron microscopy. The microbicidal effect on pathogens of clinical and agro-food interest was also evaluated. As a result it was concluded that living organisms and/or their metabolic products, can be an alternative for clean nanomaterials production with excellent antimicrobial properties.

Universidad de Pamplona, Facultad de Ciencias Básicas, Departamento de Microbiología. Grupo de Investigación en Nanotecnología y Gestión Sostenible (NANOSOST). Km. 1 Vía Bucaramanga, Pamplona Norte de Santander-Colombia. e-mail: [email protected] 137

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Introduction According to the inventory of the Project on Emerging Nanotechnologies, silver nanoparticles (AgNP), are located at the bottom of the products generated through this technology . AgNP, have received special attention because of its low volatility, high stability, long and broad antimicrobial activity [1]. At nanoscale dimension, silver presents a considerable number of applications due to their size, shape, aggregation and coupling with different molecular receptors. This characteristic facilitates using them as microbicidal agents able to release silver cations into different types of cell [2] (Figure 1). Actually, AgNP are a promising alternative to fight pathogenic organisms, which have acquired resistance to antibiotics [3] or to eliminate pathogens in different matrices [4].

Figure 1. Silver nanoparticle effect against a bacteria cell.

The main methods used to synthetize silver nanoparticles are based on physical or chemical processes, which generate waste, that in some cases are highly polluting. Therefore, there exists a need for researching about more environmental friendly methods. In the last 10 years, scientific and technological advances in 138

Biotechnological synthesis of silver nanoparticles

Nanoscience and Nanotechnology in Colombia, have permeated many higher education institutions, including the Universidad de Pamplona, Colombia. The national trend, led by the Network of Nanotechnology is to solve problems with social impacts based on the introduction of inexpensive nanoprocesses and nanomaterials [5]. The research group Nanotechnology and Sustainable Management (NANOSOST) at the Universidad de Pamplona, Norte de Santander, has focused part of their studies on the biotechnological synthesis of silver nanoparticles. The use of living organisms and /or their metabolic products seems to be a clean production strategy of high performance [6]. Filamentous fungi, are eukaryotic, with ubiquitous distribution, easy handling and nutritionally undemanding. They have the ability to become a microfactory of nanoparticles with low cost, high production efficiency and low toxicity [7]. These microorganisms secrete large amounts of bioreactives substances [8] and produce enzymes [9], which can be used as reducing agents in the biological production of AgNP. This biosynthetic capacity is mainly due to their high growth rate and high adaptability to the substrate [10]. Biotechnological synthesis of AgNP mediated by fungi can occur in two ways, “intra- or extracellularly.” In both cases, the reaction occurs thanks to the presence of substances, typically proteinaceous with catalytic activity. This reaction is produced as a defense mechanisms of the microorganisms when they are faced to metal ions, giving as a product small nanoparticles [11]. Filamentous fungi capable of synthesizing nanoparticles have shown the formation of AgNP on the mycelium (intracellular synthesis, Figure 3 A-B). However, the use of this route requires additional processes of cell disruption to separate the 139

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Figure 2. Methods used in the synthesis of Silver Nanoparticles. Advantages and disadvantages.

nanoparticles from other cellular components, thus increasing process complexity. On the contrary, extracellular synthesis allows the obtention of nanoparticles from aqueous solutions derived from the fungal biomass. A total substrate reduction without waste generation and therefore a 100% efficiency is achieved. Furthermore, colloidal solutions exhibit good dispersion with sizes ranging from 5 to 50 nm [12] (Figure 3C-3D).

Production in the laboratory Aspergillus flavus is a fungi able to grow and contaminate foods, especially cereals [13]. However, this fungus has also been studied for their ability to biosynthesize silver nanoparticles [9]. Therefore, in this research, A. flavus common pathogenic fungi found in cocoa crops was used to biosynthesize AgNPs. The process contemplated six steps, from obtaining cocoa samples to the characterization of the nanoparticles. 140


Figure 3. Biosynthetic routes of AgNP using fungi. A) Fungi biosynthetic capacity B) Mycelial after cell disruption C) Fungi growth on culture media D) Silver Nanoparticles Solution obtained by enzymatic action of fungi.

Sample collection The pods were sampled in cocoa farms from the Department Norte de Santander. Diseased pods exhibiting symptoms such as deformations, black stains, presence of yellow halos and cream colour powder. The pods in question were packed in plastic paper, labelled and transported in boxes to the laboratory for their processing.

Fungal Isolates Aspergillus flavus was obtained from ill cocoa pods exhibiting 141

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the symptoms previously mentionated. By using a mycological ase, spores containing in the upper pod cortex was directly plated on PDA. Media were incubated at 25 °C/5 days. Pure culture was carried out from the heterogeneous growth until obtaining axenic cultures, which were morphologically and molecularly characterized. Figure 4 shows the A. flavus isolating process from ill cocoa pods.

Figure 4. Sample collection and in vitro fungi isolating A) Cocoa pod ill B) Isolating process C) Plate on culture media.

Morphological Characterization Morphological characterization was performed taking into account aspects such as texture, edge, and mycelium color. Reproductive structure (spores), type of hyphae were observed by coloring them with lactophenol blue. The photographic record was obtained in a Nikon Eclipse 80i phase contrast optical microscope (100 X magnification).

Molecular Characterization DNA was isolated by using the ultraclean microbial DNA 142


isolation kit (Mo Bio Laboratories) and prepared according to their specifications. The molecular characterization of the isolated DNA was carried out by employing two molecular markers corresponding to ITS region and the gene β-tubulin. These primers are presented in Table 1. The DNA thus amplified was sequenced by MacroGen and analyzed through the BLAST database. Table 1. Primer genes β-tubulin and ITS region employed for the characterization of biocontrol fungi isolated from soil cocoa crops. Primers β-tub Primers IT S

Bt-Lev

GTC AAC TCC ATC TCG TCC ATA

BT-T 2M

CAA CTG GGC TAA GGG TCATT

PN3

CCG TTG GTG AAC CAG CGG AGG GAT C

PN16

TCC CTT TCA ACA ATT TCA CG

Biosynthesis of AgNPs by using A. flavus A. flavus was grown in 250 mL flasks containing malt broth (Oxoid) and prepared according to the specifications. The flasks were incubated in Bioshaker Plus (Molecular Technologies) at 200 rpm/25 °C in darkness. The biomass obtained was collected after 5 days (Figure 5). Subsequently, several washes were applied in order to remove residues of culture medium [9]. Then, the biomass was placed in a flask containing 100 mL of sterile distilled water to create nutritional stress to the fungus and thus obtain secondary metabolites (enzymes/reduction agents) [14]. Fungal solution obtained was filtered with Whatman No. 1 and exposed to a solution of silver nitrate (AgNO3) (Sigma-Aldrich) 1mM pH 6.5 in relation V: V, which was maintained at 25 ° C under constant stirring and darkness until change in colour (Adapted from [15]). 143

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Figure 5. Fungal Biomass Production in a Batch Bioreactor.

Evaluation of the microbial effect of AgNPs In culture media Petri dishes (60 x 15 mm) were prepared with trypticase soy agar (TSA) for bacteria and potato dextrose agar (PDA) for fungi. Each dish was plated with a different type of microorganisms. Escherichia coli were one of the tested strains. This bacterium is commonly associated with outbreaks of gastroenteritis in children [16]. Staphylococcus aureus, was also analyzed. This pathogen is frequently related with outbreaks of toxic infection through food consumption, especially contaminated grain. Finally, the fungus Candida albicans responsible for infections in the genito-urinary women tract [17] was studied. All strains were provided by the type culture center of the Universidad de Pamplona, Colombia. 144


The experimental assay consists on determining the inhibition capability of the AgNP on the tested pathogens. For that purpose, a sensidisc was placed on the center of the petri disc and a colloidal solution of the nanoparticles was added. Growth inhibition around the sensidisc was followed during 24 to 96 hours incubating at 37 ° C and 25 ° C, for bacteria and fungi, respectively. Petri dish without AgNP and inoculated with the pathogens was employed as positive control adapted from [18]. All Assays were repeated twice.

Fruit in package Raw and disinfected tomato and ground-cherry were placed inside polypropylene packaging modified with silver nanoparticles. After that, vacuum to 95% was applied (packing Brand Citalsa) and subsequently fruits were store at room temperature and cooling, with relative humidity of 45% and 55%, respectively. After three days, microbiological analysis of E. coli, aerobic mesophilic bacteria, molds and yeasts were performed [19]. Fruits without cleaning process were used as control.

Results and discussion Aspergillus was successfully isolated from diseased pods. Macroscopically, it presented a granular and olive green growth. Microscopically, conidiophores, vesicles, matulas, phialides and typical conidia of this fungal species were observed (Figure 6). Molecular characterization allowed confirming the specie A. flavus (RID 7NOVJZ7R01R-NCBI). The fungus grew favourably in batch culture and enzyme content was obtained. It was exposed to a solution of AgNO3 (0.01 M) and color change was observed after just 24 hours. Figure 7 shows the obtained nanoparticle colloidal solutions. 145

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Figure 6. Macro and Microscopic Characteristics of A. flavus.

Figure 7. Colloidal Silver Nanoparticles Solution A) enzyme solution obtained from the fungus B). Colloidal solution indicator biosynthesis Silver Nanoparticles. 146


The colorimetric result of AgNPs was checked by using UV-Vis (Spectrophotometer Shimadzu) UV-2600 with a sweep from 200 to 900 nm in a time range of 24 to 48. As negative control absorbance 0.001m AgNO3 solution was measured. The UV-Vis allows knowing the plasmon resonance profile as a product of the collective excitation of surface electron cloud present in the nanoparticles. Noble metals like silver have a plasmon resonance in the range of the visible spectrum, resulting in colloidal dispersions of nanoparticles with a range of bright colours [5]. According to several studies, the formation of AgNP can be evidenced by the formation absorbance peak located in a range between 378nm and 420nm. Values tend to increase over time due to aggregation process. Figure 8 shows the characteristic peak obtained at 420 nm, indicating the formation of AgNP using as reducing agent the metabolic products of A. flavus.

Figure 8. UV-Vis spectrum of AgNP biotechnologically synthesized using A. flavus.

The exact mechanism of formation of nanoparticles with fungi is still under study. It is believed, that substrate (usually 1 mM AgNO3) is reduced by enzymatic action of NADH dependent reductases. In addition, proteins and polysaccharides produced 147

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by fungi would be responsible of the nanoparticles stabilization. Scanning electron microscopy (Zeiss-EVOHD15) was used to determine size and shape of the biosynthesized nanoparticles. Silver nanoparticles presented spherical shapes, with sizes ranging from 10 nm to 80 nm (figure 9).

Figure 9. SEM image of silver nanoparticles.

The colloidal solution of nanoparticles was dispersed in water for 2 months in the dark (Adapted from [15]). Through this time, it was observed slights changes in colour and the formation of small aggregates. This can be explain by the principle that in a colloidal suspension, van der Waals interactions are presented causing aggregation. As a result, big clusters can appear interfering in the attachment to the cell membrane and therefore reducing the microbicidal activity. In order to obtain stable solutions, there should be electrical charges on the surface of the particles creating electrostatic repulsion and thus particles remains stable [5]. In the case of our nanoparticles, they present on their surface protein groups of fungal origin, giving stability mainly due to steric interactions. 148


In addition, physical processes like sonication by using ultra sonicates (EImasonic S 15H) were also applied before use.

Figure 10. Schematic representation of the reduced microbicidal effect of AgNP when cluster formation occurs.

The inhibitory effect of AgNP on the studied pathogens was confirmed in the petri dish experiments. The results showed that nanoparticles had had a stronger inhibitory effect on C. albicans (30,21 mm) and E. coli (28,42 mm) than S. aureus (20,78 mm) (Figure 11). This phenomenon may be to differences in cell wall composition.

Figure 11. Inhibitory effect of AgNP on the macroscopic growth of A) C. albicans B) E. coli C) S. aureus. Reproduced with permission from [19]. 149

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Once it was proved that AgNP displayed microbicidal effect on bacteria and fungi, assays were performed directly on package used for storage minimally processed fruit. It was observed a significant reduction in counts of aerobic mesophilic bacteria (AM), total coliforms (TC), fecal coliforms (CF), molds and yeasts (M and L) as a result of the use of silver nanoparticles. It can be observed in Figure 12 that nanoparticles cause a reduction from 2 to 4 log units growth in fruits cleaned and disinfected. The most significant effect of inhibiting (4 Log) was observed in fruits disinfected, stored at refrigeration temperatures and modified with AgNP, in contrast with those stored in unmodified packaging without disinfecting.

Figure 12. Inhibition effect of AgNP on storage minimally processed fruits. Reproduced with permission from [19].

Conclusions The ability of the pathogenic fungus Aspergillus flavus isolated from cocoa crops to biosynthesize silver nanoparticles was shown. The method of biotechnological synthesis allowed obtaining high amount of nanoparticles in short time with excellent phenomenological (size, distribution, stability) and microbicidal properties. These aspects allow us to expect the wide range of 150


applications of these nanomaterials, including the biocontrol of pathogenic microorganisms in different matrices. However, further investigation is required, especially regarding to potential cytotoxic effects previous scaling.

References [1] Bharathidasan, R.; Panneerselvam A. Biosynthesis and characterization of silver nanoparticles using endophytic fungi Aspergillus concius, Penicillium janthinellum and Phomosis sp. International Journal of Pharmaceutical, Sciences and Research. 2012, 3, 3163-3169. [2] Valencia, P.; Pridgen, E.; Minsoun, M.; Langer, R.; Farokhzad, O.; Karnik, R. Microfluidic Platform for Combinatorial Synthesis and Optimization of Targeted Nanoparticles for Cancer Therapy. ACS Nano. 2013, 7(12), 10671–10680. [3] Bindhu, M.R.; Umadevi, M. Antibacterial and catalytic activities of green synthesized silver nanoparticles. Spectrochimica Act Part A: Molecular and Biomolecular Spectroscopy. 2015, 135-274. [4] Villamizar, R. Nanotecnología al servicio de la conservación arquitectónica, Estrategias basadas en nanotecnología para reducir el crecimiento de hongos en ambientes abitados. Revista nano Ciencia y Tecnología. 2013, 1(1), 8-12. [5] González, E.; Puntes, V.; Casals E. NANOMATERIALES. Nanopartículas Coloidales. Nanocitec. 2015. Pag. 306. ISBN 978-958-46-6931-5. [6] www.unipamplona.edu.co/nanosost [7] Sunkar, S.; Nachiyar, C. Endophytic Fungi Mediated Extracellular Silver Nanoparticles as Effective Antibacterial Agents. International Journal of Pharmacy and Pharmaceutical Sciences. 2013, 3, 99-100. [8] Li, G.; He, D.; Qian, Y.; Guan, B.; Gao, S.; Cui, Y.; Yokoyama, K.; Li, W. Fungus- Mediated Green Synthesis of Silver Nanoparticles Using Aspergillus terreus. International Journal Meolecular Sciences. 2012, 13, 466-476. [9] Moharrer, S.; Mohammadi, B.; Azizi, R.; Yargoli, M. Biological synthesis of silver nanoparticles by Aspergillus flavus, isolated from soil af Ahar cooper mine. Indian Journal of Science and Technology. 2012, 5, 2443-2444 [10] Honary, S.; Barabadi, H.; Gharaei-Fathabad, E.; Naghibi, F. Green Synthesis

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of Silver Nanoparticles Induced by the Fungus Penicillium citrinum. Tropical Journal of Pharmaceutical Research. 2013, 12(1), 7-11. [11] Ahmad, S.R.; Minaeian, S.; Shahverdi H. R.; Jamalifar H.; Nohi A. A. Rapid synthesis of silver nanoparticles using cultura supernatants of Enterobacteria: A novel biological approach. Process Biochemestry. 2007, 42, 919-923. [12] Bhainsa, S.D. Extracellular biosynthesis of silver nanoparticles using the fungus Aspergillus fumigatus. Colloid and Surfaces. 2006, 47: 160-164. [13] Villamizar, R.; Maroto, A.; Rius F. Rapid detection of Aspergillus flavus in rice using biofunctionalized carbon nanotubes field effect transistors. Analytical and Bioanalytical Chemistry. 2010, 399:119-126. [14] Meareg, G.; Amare, P. Molecular mechanisms of Aspergillus ﬂavus secondary metabolism and development. Fungal Genetics and Biology. 2014, 66, 11-18. [15] Vigneshwaran, N.; Ashtaputre, N.M.; Varadarajan, P.V.; Nachane, R.P.; Paralikar, K.M.; Balasubramanya, R.H. Biological synthesis of silver nanoparticles using the fungus Aspergillus flavus. Materials Letters. 2007, 61, 1413–1418. [16] Foster, M. A.; Iqbal, J.; Zhang, C.; McHenry, R.; Cleveland, B.; Herazo, Y.; Fonnesbeck, H.; Payne, D.; Chappell, J.; Halasa, N.; Gómez-Duarte, O. Enteropathogenic and enteroaggregative E. coli in stools of children with acute gastroenteritis in Davidson County, Tennessee. Diagnostic Microbiology and Infectious Disease. 2015, 83, 319-324. [17] Villamizar, R.; Maroto, A.; Rius, FX. Improved detection of Candida albicans with carbon nanotube field-effect transistors. Biosensors and Bioelectronics. 2009, 136, 451-457. [18] Nasrollahi, A.; Pourshamsian, Kh.; Mansourkiaee, P. Antifungal activity of silver nanoparticles on some of fungi. International Journal of Nano Dimension. 2011, 1, 233-239. [19] Villamizar, R.; Monroy, L. Using silver nanoparticles for control pathogenic microorganisms in foods. Alimentech. 2015, 13, 54-59.

152

Ecotoxicology


Ecotoxicology in nanotechnologies Andrea Luna-Acosta

S

Several chemical compounds are used in human activities such as nanotechnologies, and are released every day in the environment. This chapter will provide an overview of sublethal and lethal effects that may have this type of compounds on living organisms and humans. It will also be devoted to ERA (Environmental Risk Assessment) and ERM (Environmental Risk Management) processes, which have been developed by environmental agencies, industries and governments in order to detect, reduce and avoid adverse effects of pollutants on ecosystems and their components. This chapter will also show how ecotoxicological tools (biomarkers, bioindicators, bioassays, biomonitoring) are very useful and necessary in this context, with some examples and case studies. There are two ways of reading this chapter. The “rapid way” consists on easily obtaining key information and methodologies of ecotoxicology for nanotechnologies, by reading only the text boxes, figures and tables. The “long way” consists on going deeper on the understanding of the concepts related with ecotoxicological research, by reading the whole chapter. Department of Ecology and Territory, Faculty of Environmental and Rural Studies (FEAR), Pontificia Universidad Javeriana, Transv. 4 No. 42-00, Bogota, Colombia. e-mail: [email protected] 155

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Eco….what? Let´s start with some definitions This section will cover definitions of the main terms that will be used in this chapter, starting with a definition for the term “ecotoxicology”. Ecotoxicology was first defined by the French toxicologist René Truhaut, in 1977 [1]. It combines two different disciplines: ecology (“the scientific study of interactions that determine the distribution and abundance of organisms”; [2]) and toxicology (“the study of injurious effects of substances on living organisms”, usually humans; [3]). Ecotoxicology aspires to assess the impact of chemicals on individuals, populations and ecosystems. It was after World War II (1939-1945) that increasing concern about the impact of toxic chemicals on the environment led toxicology to expand from studies on humans to studies on the environment. It was based on the assumption that if humans can be affected by chemical compounds (as it has been confirmed by multiple case studies), therefore animals, plants and their habitats may also be affected. In this context, the fate of pollutants in the environment is also studied in this field (Figure 1). The fate corresponds to the transport, transformation and breakdown of pollutants in the environment and within the organisms. Therefore, ecotoxicology can be defined as “the study of the harmful effects of chemicals upon ecosystems” [4]. Both terms, contaminants or pollutants, are used for chemicals that are found at levels judged to be above those that would normally be expected. However, pollutant carries the connotation of the potential to cause harm, whereas contaminant is not harmful. Nevertheless, a contaminant can become harmful and therefore, become a pollutant, if noxious effects are observed [4]. Thus, it is preferable to use the term contaminant rather than pollutant, if noxious effects have not yet been observed. The term xenobiotic 156

Ecotoxicology in nanotechnologies

Figure 1. Main goals and objectives of ecotoxicology.

is less commonly used since it has a more general sense, and corresponds to any foreign substance found within a living being [4]. More recently, the term emerging contaminant has been used for chemicals that are not commonly monitored in the environment but have the potential to enter the environment and cause known or suspected adverse effects, e.g. UV filters, pharmaceuticals, etc. Chemical compounds can be degraded by biological or chemical processes in the environment, but when degradation does not occur, they tend to accumulate, especially in sediments [5]. Factors in the environment, such as temperature, water, salinity, pH, or oxygen concentration, will determine the chemical form of chemical compounds in the environment. These factors will also determine the bioavailability of chemical compounds, which means the actual amount of substance that could exert an effect on the living organism, according to the amount that the 157

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organism has adsorbed or absorbed. Adsorbed means that the chemical enters the organism dissolved in a liquid or solid, and absorbed that it crosses an external surface to enter the organism. In aquatic ecosystems, chemical compounds can be present in water, sediments or food, and enter in that way into living organisms [6]. Inside the living organism, these compounds can be bioaccumulated (bioconcentrated or bioamplified). Bioaccumulation means an increase in the concentration of a chemical in a biological organism over time, compared to the chemical’s concentration in the environment. Bioaccumulation occurs because the chemical compound accumulates in the living being and is taken up and stored faster than it is broken down (metabolized) or excreted. Bioconcentration means bioaccumulation of a chemical in a living being when the source of chemical is solely water. Bioamplification corresponds to the increasing concentration of a chemical in the living being at successively higher levels in a food chain [6]. Hydrosoluble compounds (soluble in water) are generally less accumulated in living organisms than liposoluble compounds (soluble in lipids). However, if hydrosoluble compounds, such as many pesticides, are continuously and massively used by humans, living organisms will be constantly exposed to this type of compounds, and a constant exposure to these compounds will potentially cause chronic effects. Lipophilic compounds, such as PAHs (polycyclic aromatic hydrocarbons), PCBs (polychlorobiphenyls), PBDEs (polybromodiphenylethers) and OCPs (organochlorine pesticides), tend to accumulate more easily in living organisms. Biodisponibility and bioaccumulation of contaminants can be correlated to the appearance of noxious effects in the living being, and therefore, bioaccumulation of lipid compounds may cause major noxious effects on main physiological systems such as the endocrine, reproductive, nervous and immune systems, and may cause major ecological impacts [7]. However, it is important to 158


keep in mind that a contaminant at very high concentrations in the environment can be harmless while another contaminant at very low concentrations can be very noxious or even lethal. This can be evaluated, for example, by determining the LD50 of the compound. LD50 (abbreviation of lethal dose and 50%) is the dose of substance that is lethal for half (50%) of the animals tested. In general, the smaller the LD50 value, the more toxic the chemical is. It is expressed in milligram (mg) of toxic substance per kilogram (kg) of weight of animal, or parts per million (ppm), and it is recommended to indicate the animal that was tested (rats, rabbits, etc.). In that way, it can be extrapolated to human beings [8]. Occasionally, LD0 (lethal dose for 0% of the animals tested), LD10 (lethal dose for 10% of the animals tested) and LD90 (lethal dose for 90% of the animals tested) are also used. The term LC50 (lethal concentration for 50% of the animals tested) is also often used and is expressed in micrograms (or milligrams) of the material per liter, or ppm, of air or water. The LOAEL (lowest observed adverse effect level) can also be used and corresponds to the lowest dose of a chemical that produce a significant adverse effect [8]. Effects of contaminants vary according to extrinsic factors (chemical nature of the compound, solubility in water, persistence in the environment, chemical similarity with other molecules within the living organism), and to intrinsic factors (capacity of the living organism to transform, metabolize, absorb and/ or eliminate this compound). Inside the living organism, these substances can be transformed by detoxification processes, into less toxic molecules that are more easily eliminated. In other cases, these toxic substances are transformed by detoxification processes, into more toxic molecules than the parent molecule (Figure 2). The interaction of these molecules with DNA, proteins or steroids within the organism, may exert indirect physiological effects, especially by affecting defence mechanisms, which induces the development and/or accentuates the existence of diseases [6]. 159

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Figure 2. Responses of organisms to deleterious effects to the exposure to pollutants.

Because of these interactions within the living organism, chemical contaminants can be carcinogens, teratogens, mutagens, genotoxic substances or endocrine disruptors: • Carcinogenic substances may cause or increase the risk of cancer. • Teratogenic substances may affect the human embryo or foetus after the pregnant woman is exposed to the substance, causing physical malformations, problems in the behavioural or emotional development of the child, and decreased 160


• • •

intellectual quotient in the child, or affect pregnancies and cause complications such as preterm labours, spontaneous abortions, or miscarriages. Mutagenic substances may induce or increase the frequency of mutations in a living being above the natural background level. Genotoxic substances are mutagenic substances that specifically damage the DNA. Endocrine disruptors are substances that may interfere with the body’s endocrine system and produce adverse developmental, reproductive, neurological, and immune effects in humans, such as lowered fertility or increased female-to-male birth ratios. In wildlife, it can also induce lowered fertility or the development of imposex or intersex [9]. Imposex corresponds to the development of male sex organs in females. Intersex corresponds to the presence of both male and female cells in a single living being.

These effects can be evaluated through the measurement of bioindicators or biomarker responses. In ecotoxicology, a bioindicator is a living being whose function, population, or status can reveal the presence and/or impact of pollutants in the environment. Bioindicator species effectively indicate the condition of the environment because of their medium tolerance ranges for abiotic environmental conditions and their medium sensitivity to environmental changes [10]. Ubiquitous species have broad tolerance ranges for abiotic environmental conditions and are not very sensitive to environmental changes, while rare species have very narrow tolerances for abiotic environmental conditions and are too sensitive to environmental changes or too infrequently encountered [10]. Therefore, bioindicator species must be in between ubiquitous and rare species. Nevertheless, lethal effects of pollutants are not always observed on bioindicator species, but this does not necessarily 161

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mean that pollutants are not having a sublethal effect on living organisms. A sublethal effect do not kill but induces a stress in the organism or a stress response at cellular, biochemical and physiological levels. Both acute and chronic stress can induce a failure on physiological homeostasis (i.e. on the balance of essential physiological states), which may cause or contribute to the development of illness. In that case, it will be necessary to wait months, years, or even decades, or that the concentration of the contaminant increases significantly in the environment, to detect effects in bioindicator species. At that point, it may be very difficult to eliminate the stressor or the effect that it has on the environment. Therefore, when effects are not observed at the level of the individual or the population, the effects of contaminants can be evaluated through the measurement of biomarker responses, at a smaller, more sensitive and shorter time scale levels. This may help considerably to prevent irreparable environmental damages. In medicine, the National Institutes of Health Biomarkers Definitions Working Group defines a biomarker in human beings as “a characteristic that is objectively measured and evaluated as an indicator of normal biological processes, pathogenic processes, or pharmacologic responses to a therapeutic intervention”. PSA (prostate-specific antigen) is an example of biomarker used in medicine as a proxy of prostate size with rapid changes potentially indicating cancer, allowing early diagnosis of dysfunctions on human health. Because of the utility of biomarkers as predictive tools, their scope has extended to other areas such as ecotoxicology. In ecotoxicology, biomarkers allow early diagnosis of dysfunctions on living organisms, populations and ecosystems, before substantial damage occurs in the environment [11]. They correspond to biological responses in living organisms at molecular, biochemical, cellular, physiological or behavioural levels that allow to highlight and demonstrate the exposure and/or effects of contaminants 162


[11,12]. Examples of biomarkers range from the inhibition of AChE (acetylcholinesterase) in the nervous system, an enzyme that if it is inhibited, it may induce changes in the behaviour of living organisms, to the thinning of eggshells in birds. Biomarkers can help to bridge the gap between the laboratory and the field by giving direct evidence of whether or not a particular animal, plant or ecosystem is being affected by pollution. They will often provide more reliable evidence of exposure to pollutants than measurements of the pollutants themselves in the environment, because pollutants are often short-lived and difficult to detect, whereas their effects (detected through biomarker responses) may be much longer-term. Different extrinsic factors or stressors (climate, physicochemical variables, etc.) in the environment, including chemical contaminants, may exert a stress in the living organisms, and therefore in the environment. According to Van Straalen (2013), “a situation of stress arises when some environmental factor changes and an organism finds itself outside its ecological niche. By definition, the organism cannot grow and reproduce outside its niche, but it may survive temporarily. The stress can be relieved by moving back to the niche, [by using behavioural mechanisms or suppressing the stressor, and during that time by developing a temporary physiological adaptation, which allows survival until the stressor is gone], or by changing the boundaries of the niche (genetic adaptation) [13]”. In this context, responses of biomarkers of environmental stress can be measured and will allow to early assess ecosystem health, and even more if these responses are measured in keystone species. A keystone species is a species that if it is removed from the environment, it will have a disproportionately large effect on its environment relative to its abundance, which means that it will generate a dramatic shift in the ecosystem, affecting many other species, even though that species was a small part of the ecosystem [14]. 163

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These effects can be observed by carrying out bioassays, which correspond to scientific experiments that use a live animal or plant (in vivo) or tissue or cell (in vitro) to determine the effects of a substance on a living organism. In some cases, bioassays will try to “reproduce” natural environmental conditions and this can be done at different scales: • In microcosms: at laboratory scale model ecosystem. • In mesocosms: in an enclosed body of water (pond or flow through system) of close to natural conditions for running controlled aquatic experimentations. • In macrocosms: in large model systems. Active biomonitoring is in between bioassay and biomonitoring. It corresponds to transplantation experiments in the field, in which animals from a clean site are transferred to a contaminated site and vice-versa for long periods of time in order to evaluate their responses in a clean and/or contaminated environment. This type of monitoring is usually combined with chemical monitoring, which corresponds to the assessment of concentrations of chemical compounds (potential contaminants) in water or sediments [6]. Biomarker responses can be measured in bioassays, active biomonitoring and biomonitoring, and are very useful for ERA (Environmental Risk Assessment) and ERM (Environmental Risk Management) processes [15]. ERA is the process that estimates the magnitude and probability of adverse effects of pollutants and other anthropogenic activities on ecosystems and their components, while EIA (Environmental Impact Assessment) is the process that consists on estimating the magnitude of impacts of chemical contaminants in the environment [12]. An assessment is scientifically oriented and consists on evaluating and describing the impacts (with the analysis of data using quantitative techniques and scientific methodologies), while 164


management is more politically oriented and consists on choosing among alternatives and determining acceptability of risks (with the examination of solutions to the problem and the establishment and application of regulatory measures, laws and public policies). In that sens ERM is defined as the process of choosing among alternatives and determining acceptability of risks of chemical contaminants in the environment. In this chapter we will consider only ERA and ERM, which should be done before any impact is detected and therefore before carrying out EIAs (Figure 3).

Figure 3. ERA (Environmental Risk Assessment), ERM (Environmental Risk Management) and ecotoxicological tools.

Chemical Contaminants Several chemical compounds are used on a daily basis in human activities, such as agricultural, farming, mining, industrial and domestic activities. Some examples of these chemical compounds are heavy metal particles including NPs (nanoparticles). These 165

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compounds are not necessarily immediately toxic for living organisms, but their effect can be sublethal and chronic. They are considered as pollutants in the environment because they have shown to be carcinogenic, teratogenic, mutagenic, genotoxic or endocrine disruptors.

Heavy metals and nanoparticles As (arsenic), Ba (barium), Be (beryllium), Cd (cadmium), Cr (chromium), Cu (copper), Hg (mercury), Mn (manganese), Ni (nickel), Pb (lead), Se (selenium), Sn (tin), Tl (thallium) and Zn (zinc) are some of the metals called ‘heavy’ because of their high relative atomic mass. These metals persist in nature and can cause damage or death in animals, humans, and plants, even at very low concentrations. For example, the maximum consumption limit recommended for Cd is 5 µg/l in water and 1 µg/kg/day in food, according to international organizations, such as EPA (Environmental Protection Agency), FAO (Food and Agriculture Organization) and WHO (World Health Organization) [16]. All these compounds are natural components of the Earth´s crust. They can enter a water supply by industrial and consumer waste, or even from acidic rain, breaking down soils and releasing heavy metals into streams, lakes, rivers, and groundwater. As trace elements, which means in very small quantities, some heavy metals (e.g. Cu, Se and Zn) are essential to maintain the metabolism of the human body. However, at higher concentrations they can lead to poisoning. Poisoning can result, for instance, from drinking-water contamination (e.g. lead pipes), high ambient air concentrations near emission sources, or intake through the food chain. Other heavy metals (e.g. As, Cd, Hg, Pb) are not essential to maintain the metabolism of the human body and can be toxic or exert noxious effects at low concentrations. Safety levels in food or water have been prescribed for some of them by international organizations 166


(Table 1). Unfortunately, an important number of pollution accidents and disasters in the world are due to the inappropriate Table 1. Principal metals used in nanotechnologies and quality standards for humans, to protect public health. Own adaptation based on different international sources. Regulations and recommendations can be expressed as “not-to-exceed” levels, that is, levels of a toxic substance in air, water, soil (not shown in this table), or food that do not exceed a critical value that is usually based on levels that affect animals; they are then adjusted to levels that will help to protect humans. Sometimes these not-to-exceed levels differ among federal organizations because they use different exposure times, animal studies or other factors. EPA: Environmental Protection Agency; FAO: Food and Agriculture Organization; WHO: World Health Organization; OSHA: Occupational Safety and Health Administration; ww: wet weight; ND: standard not determined. Institution

EPA

FAO/ WHO

FAO/ WHO

OSHA

WHO

Source

Drinking water

Food grade

Fish

Air£

Humans (organ or tissue)

mg/l (ppm)

mg/ kg ww (ppm)

mg/kg ww (ppm)

mg/m3

Longest half-life (days)

Metals

Units

Ag (Silver)

0.100

ND

ND

0.010

50

As (Arsenic)

0.010

0.5

50.0

0.010

3

Au (Gold)

ND

ND

ND

ND

ND

Ba (Barium)

2.000

ND

ND

0.500