Wetlands ecology, conservation and restoration

WETLANDS: ECOLOGY, CONSERVATION AND RESTORATION No part of this digital document may be reproduced, stored in a retrieval system or transmitted in any form or by any means. The publisher has taken reasonable care in the preparation of this digital document, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained herein. This digital document is sold with the clear understanding that the publisher is not engaged in rendering legal, medical or any other professional services.

WETLANDS: ECOLOGY, CONSERVATION AND RESTORATION RAYMUNDO E. RUSSO EDITOR Nova Science Publishers, Inc. New York

Copyright 2008 by Nova Science Publishers, Inc. All rights reserved. No part of this book may be reproduced, stored in a retrieval system or transmitted in any form or by any means: electronic, electrostatic, magnetic, tape, mechanical photocopying, recording or otherwise without the written permission of the Publisher. For permission to use material from this book please contact us: Telephone 631-231-7269; Fax 631-231-8175 Web Site: http://www.novapublishers.com NOTICE TO THE READER The Publisher has taken reasonable care in the preparation of this book, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained in this book. The Publisher shall not be liable for any special, consequential, or exemplary damages resulting, in whole or in part, from the readers use of, or reliance upon, this material. Independent verification should be sought for any data, advice or recommendations contained in this book. In addition, no responsibility is assumed by the publisher for any injury and/or damage to persons or property arising from any methods, products, instructions, ideas or otherwise contained in this publication. This publication is designed to provide accurate and authoritative information with regard to the subject matter covered herein. It is sold with the clear understanding that the Publisher is not engaged in rendering legal or any other professional services. If legal or any other expert assistance is required, the services of a competent person should be sought. FROM A DECLARATION OF PARTICIPANTS JOINTLY ADOPTED BY A COMMITTEE OF THE AMERICAN BAR ASSOCIATION AND A COMMITTEE OF PUBLISHERS. LIBRARY OF CONGRESS CATALOGING-IN-PUBLICATION DATA Wetlands : ecology, conservation, and restoration / Raymundo E. Russo (editor). p. cm. ISBN 978-1-60876-354-2 (E-Book) 1. Wetland ecology. 2. Wetland conservation. 3. Wetland restoration. I. Russo, Raymundo E. QH541.5.M3W4836 2008 577.68--dc22 2008030635 Published by Nova Science Publishers, Inc. New York

CONTENTS Preface vii Expert Commentary Two Alternative Modes for Diffuse Pollution Control by Wetlands 1 Chen Qingfeng, Shan Baoqing and Ma Junjian Short Communication Multiangular Imaging of Wetlands in New England 7 Lesley-Ann L. Dupigny-Giroux and Eden Furtak-Cole Research and Review Articles Chapter 1 Wetlands: Water Living Filters? 15 Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto Chapter 2 Remote Sensing Data for Regional Wetland Mapping in the United States: Trends and Future Prospects 73 Megan W. Lang and Greg W. McCarty Chapter 3 Transforming Useless Swamps into Valuable Wetlands: Evaluating Americas Policy, 1970-2008 113 Andrea K.Gerlak and Jeanne N. Clarke Chapter 4 Dynamics of Coastal Wetlands and Land Use Changes in the Watershed: Implications for the Biodiversity 133 Miguel ngel Esteve, M. Francisca Carreo, Francisco Robledano, Julia Martnez-Fernndez and Jess Miano Chapter 5 Pathogen Removal in Constructed Wetlands 177 Kela P. Weber and Raymond L. Legge Chapter 6 The Role of Harvest and Plant Decomposition in Constructed Wetlands 213 Juan A. lvarez and Eloy Bcares

Contentsvi Chapter 7 Nutrition and Toxicity of Inorganic Substances from Wastewater in Constructed Wetlands 247 Zhenhua Zhang, Zed Rengel and Kathy Meney Chapter 8 A Conceptual and Methodological Framework for the Study of Vegetated Fluvial Landscape Evolutionary Trajectories 271 Dov Corenblit, Johannes Steiger, Eric Tabacchi and Angela M. Gurnell Chapter 9 Macrophyte Morphological Response to the Industrial Effluent Toxicity in a Constructed Wetland 295 H. R. Hadad, M. M. Mufarrege, M. Pinciroli, G. Di Luca, V. del Sastre and M. A. Maine Chapter 10 Phytoremediation Processes for Water and Air Pollution Control in the Aspects of Nutrient and Carbon Dioxide Removals 325 Jae Seong Rhee, Yonghui Song, Fasheng Li and Janjit Iamchaturapatr Chapter 11 Phytoplankton Biomass Regulation in Contrasting Environmental States of Temporary Pools 359 Silvia Martn, Marta Rodrguez and David G. Angeler Chapter 12 Can Tern Migrants Coexist with Urban Development and Estuarine Recreational Activities? 373 Ken Chan, Jill Dening and Marja-Leena Malinen Chapter 13 Agricultural Wetlands 391 R. Krger Chapter 14 Profiling Cover Cycle Dynamics for Prairie Pothole Wetland Landscapes 407 Rebecca L. Phillips and Ofer Beeri Index 419

PREFACE Wetlands are lands where saturation with water is the dominant factor determining the nature of soil development and the types of plant and animal communities living in the soil and on its surface. Wetlands vary widely because of regional and local differences in soils, topography, climate, hydrology, water chemistry, vegetation, and other factors, including human disturbance. Indeed, wetlands are found from the tundra to the tropics and on every continent except Antarctica. This new book brings together the latest research in the field. Short Communication - Multiple view angles (MVA) or multiangular imaging represents a yet to be explored use of the remote sensing of wetlands. The ability to view the landscape off-nadir (traditionally the surface is viewed at right angles) allows for the quantification of moisture stress, species separation and the proportion of vegetation to standing water in these ecosystems. This commentary will focus on the ratio of two broadband wavelengths (near- infrared to blue) derived from multiangular images acquired by the Airborne Multi-angle Imaging SpectroRadiometer (AirMISR) of wetlands across New England. The resulting insights into the photointerpretation, monitoring and mapping of wetlands will be highlighted. Chapter 1 - Human societies have indirectly used natural wetlands as wastewater discharge sites for many centuries. Observations of the wastewater depuration capacity of natural wetlands have led to a greater understanding of the potential of these ecosystems for pollutant assimilation and have stimulated the development of artificial wetlands systems for treatment of wastewaters from a variety of sources. Constructed wetlands, in contrast to natural wetlands, are human-made systems that are designed, built and operated to emulate wetlands or functions of natural wetlands for human desires or needs. Constructed wetlands have recently received considerable attention as low cost, efficient means to clean-up not only municipal wastewaters but also point and non-point wastewaters, such as acid mine drainage, agricultural effluents, landfill leachates, petrochemicals, as well as industrial effluents. Currently, untreated wastewater discharge in the natural wetlands sites is becoming an increasingly abandoned practice whereas the use of constructed wetlands for treatment of wastewater is an emerging technology worldwide. However, natural wetlands still play an important role in the improvement of water quality as they act as buffer zones surrounding water bodies and as a polishing stage for the effluents from conventional municipal wastewater treatment plants, before they reach the receiving water streams. In fact, one of the emerging issues in environmental science has been the inefficiency of wastewater treatment plants to remove several xenobiotic organic compounds such as pesticides and pharmaceutical residues and consequent contamination of the receiving water bodies. Recent

Raymundo E. Russoviii studies have shown that wetlands systems were able to efficiently remove many of these compounds, thus reaffirming the importance of the role which can be played by wetlands in water quality preservation. The aim of this work is to present a review on the application of wetlands as living filters for water purification. Emphasis was focused on the removal of micropollutants, especially xenobiotic organic compounds such as pharmaceuticals residues, which are not efficiently removed by conventional municipal wastewater treatment plants. Furthermore, the role of wetlands as protection zones which contribute to the improvement of the aquatic ecosystems quality will be discussed. Chapter 2 - Historically, the biologic, aesthetic, and economic values of wetlands were largely unappreciated. Wetlands within the United States have been and are continuing to disappear rapidly. Efforts are being made to conserve remaining wetlands and many regulatory policies have been adopted in support of this goal. To regulate the loss, preservation, and/or restoration of wetlands and to judge the effectiveness of these regulatory efforts in preserving associated ecosystem services, wetlands must be routinely monitored. Wetland mapping is an essential part of this monitoring program and much effort has been made by the US state and federal governments, as well as other organizations, to provide quality wetland map products. Wetland maps can serve a variety of purposes including regulation and natural resource management. They can also be used to parameterize models that quantify water quality and quantity, as well as the provision of wetland ecosystem services, at the watershed scale. Wetland hydrology is the most important abiotic factor controlling ecosystem function and extent, and it should therefore be a vital part of any wetland mapping or monitoring program. New approaches are needed to not only map wetlands, but also to monitor wetland hydrology as it varies in response to weather, vegetation phenology, surrounding landuse change, and other anthropogenic forces including climate change. Recently developed remote sensing technologies and techniques have the potential to improve the detail and reliability of wetland maps and the ability to monitor important parameters such as hydrology. Various types of remotely sensed data (e.g., aerial photographs, multispectral, hyperspectral, passive microwave, radar, and lidar) have different capabilities with specific advantages and disadvantages for wetland mapping at the regional scale. Although aerial photographs were traditionally used to map wetlands and infer hydrology, fine-resolution optical images are now available more frequently as commercial agencies increase satellite coverage (e.g., Quickbird and IKONOS). However, optical data, such as aerial photographs and multispectral satellite images have limitations, including their inability to detect hydrology below dense vegetative canopies and their limited ability to detect variations in hydrology (i.e., inundation and soil moisture). The restrictions of optical data are increasingly being compensated for with the use of new technologies, including synthetic aperture radar, lidar, and geospatial modeling. The availability of these new data sources is increasing rapidly. For example, many states in the US are now collecting synoptic state-wide coverages of lidar data. The sources, strengths, and limitations of different types of remotely sensed data are reviewed in this chapter, as well as the importance of temporal and spatial resolution necessary for regional scale wetland mapping efforts. The potential of multi-temporal, multi-sensor approaches that capitalize on geospatial modeling are emphasized for meeting current wetland mapping challenges. Chapter 3 - This paper traces the evolution of Americas wetland policy beginning with passage of the Clean Water Act (CWA) of 1972. This law, for the first time, established a

Preface ix federal program to protect wetlands, dramatically elevating the value of these ecosystems. However, despite attitudinal changes and new governmental programs, the nation continues to lose its potentially valuable wetlands -- albeit at a slower rate than was the case in the 1970s and prior to the passage of the CWA. This chapter offers an objective evaluation of the federal wetlands protection policy. The authors place this evaluation within a broad societal context, showing that since 1970 there have occurred sweeping demographic, economic, and political changes that clearly have impacted the extent of wetlands in the United States. They argue that Section 404 has failed to reverse the net loss of wetlands in the U.S. Moreover, it has evolved into a policy lightening rod within the water resources arena and been a major factor in Congress failure to revise and reauthorize the Clean Water Act. Finally, the authors offer some recommendations designed to improve the policy, arguing for heightened wetlands protection through partnerships and acquisitions. Chapter 4 - The Mediterranean coastal landscapes have suffered significant changes along the last decades due to the agricultural intensification and tourist development. Such changes have modified the water flows and specifically the hydrological regime of wetlands, as has occurred in the Mar Menor (Southeast Spain). The Mar Menor coastal lagoon and associated wetlands present noticeable ecological and biodiversity values. However, the landuse changes in the watershed and the consequent changes in the water and nutrient flows along the period 1980-2005 are threatening the conservation of these wetlands. A dynamic model has been developed to simulate the key environmental and socio-economic factors driving the export of nutrients to the Mar Menor lagoon and associated wetlands, where some eutrophication processes have appeared. In this chapter the changes in the vegetal and faunistic assemblages are analysed. Vegetal communities are studied by means of remote sensing techniques, which have provided information about the changes in area and habitat composition of the wetlands along the considered period. This has shown that the habitats more negatively affected by the hydrological changes are those most threatened in the international context and with a highest interest from the point of view of biodiversity conservation. It has also been possible to verify the direct relationships between all these changes at wetlands scale and the agricultural changes at the watershed scale. Two faunistic communities especially sensitive to these ecosystemic changes have also been studied: i) Wandering beetles and ii) Birds (waterbirds and steppe passerines). Wandering beetles (Coleoptera) were studied with pitfall traps in 1984, 1992 and 2003 and steppe passeriforms with line transects in several years along the period. In both communities evident changes have been observed. Regarding beetles, the most halophilous species have been favoured, some of them especially relevant due to its rarity in the European context. The ratio Carabidae/Tenebrionidae has shown to be a good indicator of the hydrological changes of the wetlands. Waterbirds have shown dramatic changes in their relative abundances within the lagoon, with a long-term decline in the most characteristic original species, increases in generalist piscivores and a recent appearance and rapid growth of the herbivores guild. In the case of steppe passeriforms, this community has been negatively affected, especially some species like Melanocorypha calandra. The family Alaudidae has lost importance to the benefit of the families Turdidae and Fringillidae. These changes can be considered a loss of value in relation with the original passeriform community, since the wetland qualifies as a Specially Protected Area under the EUs Bird Directive, precisely on the basis of its genuine steppe bird assemblage.

Raymundo E. Russox In conclusion, the changes at wetlands scale clearly reflect the hydrological modifications at the watershed scale and have significant effects on the most characteristic biodiversity of the wetlands of coastal arid systems. Chapter 5 - Conventional secondary and tertiary wastewater treatment methods include activated sludge, trickling filters, slow sand filtration, chlorination, ozonation and UV radiation. Chlorination being the most widely used pathogen disinfection method is presently under scrutiny as chlorination can produce carcinogenic trihalomethanes when natural organic matter is present in the wastewater. Constructed wetlands (CWs) have proven to be an effective treatment alternative for the removal and inactivation of pathogens in wastewaters. Constructed wetlands have low principle and operating costs and are fairly simple to design and implement, making them an attractive wastewater treatment alternative when compared to conventional secondary or tertiary treatment processes. Constructed wetlands designed for pathogen treatment are most often preceded by filtration or sedimentation. Pathogen removal efficiencies upwards of 99.99% have been reported by multiple authors employing many different constructed wetland designs. Constructed wetland design tends to be based largely on rule of thumb sizing, as the specific mechanisms and fundamental variables involved in pathogen removal are only vaguely understood. Suggested mechanisms of pathogen treatment in CWs include but are not restricted to sedimentation, natural die-off, temperature, oxidation, predation, unfavourable water chemistry, biofilm interaction, mechanical filtration, exposure to biocides and UV radiation. Pathogen removal has been shown to correlate well with hydraulic retention time. Use of first order decay kinetics is the preferred method to describe and predict pathogen removal in CWs. A severe lack of attention has been given to the comparative quantification of the specific mechanisms contributing to pathogen treatment in constructed wetlands. Small-scale controllable constructed wetland systems are identified as systems which can be used in conducting well-designed controlled experiments where fundamental mechanisms and variables involved in pathogen removal can be comparatively quantified. It is proposed that if the fundamental mechanisms and variables affecting pathogen removal in constructed wetlands are better understood and quantified the large performance variations reported for similarly designed treatment wetland systems can be better explained, engineered and controlled. Chapter 6 - Upon decomposition, at the end of the summer and during the autumn, wetland vegetation releases organic carbon into the wetland system. A part of this organic matter remains in the wetland, and is degraded at different rates during the rest of the year. Therefore, litter decomposition has important consequences on constructed wetlands because it is related to the autochthonous production of organic matter, clogging rates in surface-flow wetlands, and terrestrialization in free-water surface wetlands. The effect of harvest was studied in two free-water surface-flow wetlands. Both wetlands were planted with Typha latifolia with one of the wetlands harvested. On the other hand, decomposition rates of Typha latifolia were quantified during both winter and summer in the non-harvested surface constructed wetland using the litter bag technique. Nutrient concentrations were always lower in the effluent of the harvested wetland, indicating nitrogen and phosphorus release by decomposition of vegetation, in the non-harvested system. In addition, harvesting reduced the effluent TSS and BOD concentrations by 37.3% and 49.2%, respectively, when compared to the non-harvested wetland in spring. Seasonal background concentrations (C*) in the wetlands, increased from winter to spring and decreased again in summer. Organic load and nutrients produced per gram of Typha were evaluated by using in-

Preface xi situ Typha degradation experiments. Taking into account the experiments of litter bag technique, no significant differences were found in both variables among the different mesh sizes, with the exception of the control bags in winter. Meso or macrofauna did not play any role in plant decomposition. Decomposition rates were significantly different between winter and summer when considering each mesh size separately. Decomposition rates from adjusted exponential models ranged from 0.0014 to 0.0026 d-1 in winter (5C), and from 0.0043 to 0.0052 d-1 in summer (20C). Typha decomposition rates were compared with others macrophytes. From these decomposition rates, it is estimated that 31% of the initial mass of plant detritus would remain in the system after one year. Based on the research conducted during several experiments, harvesting can be recommended as an operational and management strategy in warm climates and diluted wastewater conditions. Chapter 7 - The use of constructed wetlands for purification of wastewater has received increasing attention around the world. A variety of wetland plant species (including ornamental ones) as either a monoculture or species mixes are used in constructed wetlands. Plants play an extremely important role in removing pollutants from wastewater. Although there is considerable information on plant productivity, biomass and nutrient dynamics in natural and fertilized wetlands, most studies on constructed wetlands for treatment of wastewaters have only addressed general aspects of plant growth and nutrient accumulation. Nutrition and toxicity of inorganic substances such as nitrogen, sulphur, salts and metals in wastewater on wetland plants has not been fully investigated and their interactive effects and environmental cycling in constructed wetlands remain poorly understood. Nitrogen nutrition is the most important factor influencing plant performance in constructed wetlands, but higher NH4-N may become toxic to wetland plants. Sulphur is an essential nutrient for plant growth, but under waterlogged conditions sulphate is reduced to hydrogen sulphide that is highly toxic to wetland plants. Many metals in wastewater are essential micronutrients for wetland plants, but become toxic if their concentration exceeds a specific critical point. A proper amount of salts is essential for plant growth, but high concentrations of salts, particularly sodium chloride in wastewater have harmful effects on plant growth. Wetland plant species have differential capacity to take up nutrients, different preference for nitrogen forms and have evolved various adaptive mechanisms protecting them against toxicity of inorganic substances. Given that plants are an integral part of constructed wetlands, the selection of suitable species, improvement of cultivations and determination of factors affecting growth are needed to produce healthy and effective wetland ecosystems. Understanding biogeochemical cycling in wetlands as well as nutrition and toxicity of inorganic substances from wastewater on plant development and function may help reduce performance variability and enhance pollutant removal in constructed wetlands. Chapter 8 - This chapter presents a conceptual and methodological framework to study temporal and spatial changes of fluvial landforms and associated plant communities and to identify the underlying causes of either progressive or sudden changes. Mutual interactions and feedbacks between hydrogeomorphic processes, fluvial landforms and vegetation dynamics are considered within this framework, leading to the analysis of biogeomorphic (i.e., landforms and associated vegetation communities) evolution trajectories within the fluvial corridor and to the evaluation of their consequences for ecological and geomorphic forms and processes.

Raymundo E. Russoxii First, fundamental aspects linked to the conceptual model of Fluvial Biogeomorphic Succession (FBS model) proposed by the authors (cf. Corenblit et al. 2007) are presented. This model describes the most dominant biogeomorphic succession trajectory of temperate rivers under current bioclimatic and anthropogenic conditions, starting from the rejuvenated state (bare sediment within the channel after a destructive flood). This dynamic model involves a characteristic sequence of four biogeomorphic phases where interactions of hydrogeomorphic processes and vegetation dynamics are either strong or weak according to different spatiotemporal configurations. The characteristic evolutionary trajectory corresponds to a progressive shift from the dominance of allogenic (hydrogeomorphic) processes to the dominance of autogenic (ecological) processes. It is marked by a development of specific stabilised vegetated landforms such as banks, islands and floodplains. In particular, the cyclic dynamics of the biogeomorphic succession (i.e., frequency and magnitude of rejuvenation and maturation processes), incorporating critical thresholds are discussed. Second, a conceptual tool for the description and analysis of potential fluvial landscape evolutionary trajectories is proposed. This conceptual tool is a discrete three dimensional biogeomorphic phase-space composed of five key-stages of vegetation development (bare sediment; seedlings and saplings; adult herbs; adult shrubs; adult trees) within four distinct zones of the river corridor, exposed to four distinct levels of hydrogeomorphic disturbance (permanent submerged area; high flood-frequency area; low flood-frequency floodplain; non- submersible area). The four main processes controlling shifts between biogeomorphic configurations within the phase-space are related to the critical role of pioneer vegetation within fluvial landscape dynamics. Finally, a methodological basis to test and to refine the model using a probabilistic transition analysis combining the biogeomorphic phase-space, empirical field data, GIS and remote sensing at local and regional scales is proposed and its applications for river management are discussed. Chapter 9 This chapter describes the morphological variations of floating and rooted macrophytes growing in a wetland constructed for the treatment of industrial wastewater and in natural wetlands of the Middle Paran River oodplain, Argentina. Cross-sectional areas (CSA) of the root, stele and of metaxylem vessels and the total metaxylem CSA were measured. In addition, parameters such as dry biomass, chlorophyll concentration, and metal (Cr, Ni and Zn) and nutrient (P) concentrations were compared. During the first months of operation of the constructed wetland, only sewage was poured and floating macrophytes were dominant. After five years of operation, Typha domingensis was the dominant species in the constructed wetland. In this species, biomass and height of the plants at the inlet and outlet were significantly higher than in the natural wetlands. The plants growing at the inlet showed root and stele CSA values significantly higher than those for the plants growing at the outlet and in natural wetlands. The total metaxylem vessels CSA of the inlet plants were significantly higher than those obtained in the outlet and natural wetlands owing to the plants of this site showed the highest number of metaxylem vessels. In order to determine the morphological changes as an adaptive response to the contaminants present in the effluent, greenhouse experiments were carried out with P. stratiotes and E. crassipes. In P. stratiotes, Ni and Cr+Ni+Zn treatments were the most toxic ones, in which biomass, chlorophyll and the internal morphological parameters of roots decreased significantly, while in E. crassipes Ni caused toxic effects in the internal as well as the external morphology. The modifications

Preface xiii recorded account for the adaptability of T. domingensis to the conditions prevailing in the constructed wetland, which allowed it to become the dominant species. This chapter may contribute to the design and mainteinance of constructed wetlands that include the macrophytes studied. Chapter 10 - The growth of industries and major agricultural enterprises (especially food industries) supplying the human demands for their increasing population causes an annihilation of water ecosystems and an augmentation of water pollutions. These are the main sources of nutrient supplements in water resources. Excess nutrients led to the eutrophication phenomena and in many cases the deterioration of public health. While the role of carbon dioxide (CO2) gas in global climate change has become well-known, which is one of the most important environmental issues of our day, therefore it is necessary to develop technologies for the minimization of CO2 discharging into the atmosphere. Although CO2 occurs naturally in the atmosphere, its current atmospheric concentrations have been greatly affected by human activities. One ecological method used for treating polluted water containing high nutrients and encouraging CO2 sequestration is treatment wetlands, where various aquatic plants are used for purifying the water and wastewater from excess nutrients and also withdrawing the anthropogenic CO2 from polluted atmosphere into plants biomass by photosynthesis process. Although wetland area around the world has diminished and continues to lose due to economic development, agriculture, and other landscape alterations, recently many of these losses are compensated by construction of new wetlands due to an our increasing understanding of wetland functions and values on global environment. Chapter 11 - Although abiotic forces play a fundamental role in community and process regulation of disturbed wetland ecosystems, biotic interaction is increasingly recognised for having important regulatory feedback effects. This chapter reports on the context-specific role of biotic and abiotic regulation of phytoplankton biomass in temporary ponds. Contamination of artificial ponds with different application concentrations of a fire retardant resulted in alterations of the trophic status, primary producer and zooplankton communities in treatment ponds. Principal component analyses suggested that facilitation of phytoplankton biomass through cladocerans was the most important controlling factor in nutrient-limited control ponds. These biotic interaction effects disappeared in retardant treatment ponds where phytoplankton biomass was almost exclusively controlled by water depth fluctuation. This context-specific, eutrophication-mediated physical control of algal biomass in treatment ponds adds a new dimension to the traditional perspective of resource and consumer control of phytoplankton in alternative ecosystem states in lakes. The context-dependent interplay of physical and biotic processes in wetlands will likely influence applied issues and challenge wetland management and restoration. Chapter 12 - Urbanisation and recreational activities are two of the major causes of population declines of species, and throughout the world they continue to spread and intensify at a rapid rate. The two are often linkedan increase in recreational activities is often associated with nearby growth in residential development and vice versa. Developmental growth is greatest in places of high tourism value, such as in coastal areas with sandy shores. Sandy coasts are popular with beach walking and jogging, swimming, off-road vehicles, boating, ecotourism, and other outdoor activities. The most concentrated activities are in estuaries with sandbanks and intertidal flats that are protected from the open ocean. Yet the same estuaries are often sensitive ecosystems, commonly frequented by a variety of resident

Raymundo E. Russoxiv and migrant birds that use the areas to breed, forage, or roost. Increasing incidents of human disturbance can affect breeding behavior, feeding patterns, opportunities for rest, and decline in estuarine bird abundance. The direct impact on reproduction in breeding birds is obvious, but survival of migratory species is also affected through ineffective build-up of requisite fat reserves to successfully undertake their migratory journey. For both resident and migrant birds, disturbance could result in reduced feeding time, lowering the necessary fat reserves for survival. Chapter 13 - Increased agricultural production, land drainage and resultant land use changes have increased loads of non-point source pollutants being discharged into aquatic ecosystems. Estimates suggest that non-point source pollution (NPS) contributes over 65% of the total pollution load to inland surface waters, including 332,000 km of rivers, 215,000 ha of lakes and 1.5 x 106 ha of estuaries. There are two types of agricultural wetlands that could mitigate NPS pollution: constructed wetlands and surface drainage ditches. Constructed wetlands are commonly used to mitigate increased nutrient, biological oxygen demand, and pesticide loads prior to entering receiving waters. However, some farmers will forgo the practice of constructing a wetland for routing water because of associated costs of construction, maintenance and loss of land in agricultural production. Agricultural drainage ditches are management tools put in place by farmers to rapidly remove standing water from their farmland. Drainage ditch function is simply one of drainage; however, research has shown that surface vegetated drainage ditches are primary intercept wetlands characterized by an ephemerally inundated hydroperiod, developed hydro-soils and a suite of facultative hydrophytes. Studies in the mid-South US have shown vegetated surface drainage ditches to reduce both pesticide and nutrients loads within the ditch prior to effluent reaching receiving waters. This is increasingly important in todays landscape where fertilizer and pesticide applications are still high. Pollutant reduction capacity within ditches may be improved with temporal and spatial manipulation of water residence at critical junctions of non-point pollutant loss throughout the year. Primary interception, transformation and mitigation of agricultural pollutants has far reaching consequences for aquatic ecosystem health, downstream eutrophication, and coastal dynamics such as hypoxia, commercial fisheries and economic development. Chapter 14 - Over 3 million wetlands populate the U.S. portion of the Prairie Pothole Region (PPR), where conservation goals include restoration and preservation of the cover cycle. The cover cycle is characterized by seasonal and annual changes in vegetation and open water and is closely coupled to climate and natural ecosystem functions. A complete cover cycle include periods of time when high waters drown hydric vegetation during deluge and periods where hydric vegetation expands as waters dry-down during drought. Changes in wetland cover may occur on weekly, monthly, or annual time-scales. These dynamics contribute to a rich diversity of habitats that support more waterfowl than any other region in North America. In addition temporal dynamics, PPR wetlands rarely function as single entities because of shared surface and/or groundwater hydrology. This spatial interdependence requires PPR wetland functional assessments represent populations of wetlands, commonly referred to as profiles. Synoptic data profiling cover cycle stage and return time for populations of wetlands would scaffold large-scale investigations of ecosystems services, habitat status, and sensitivity to climate change. This chapter describes application of previously developed tools for synoptic delineation of wetland water and hydric vegetation cover to classify cover cycle for thousands of wetland

Preface xv basins within a single satellite image (10,000-30,000 km2 of land area). Using satellite data layers in geographic information systems (GIS), wetland profiles developed using current (2007) wetland cover data are compared with profiles developed using National Wetland Inventory (NWI) data from 1980. Results underscore the dynamic nature of these ecosystems and the need for current observations when setting conservation goals, monitoring restoration effectiveness, and evaluating anthropogenic impacts.

In: Wetlands: Ecology, Conservation and Restoration ISBN: 978-1-60456-995-7 Editor: Raymundo E. Russo 2008 Nova Science Publishers, Inc. Expert Commentary TWO ALTERNATIVE MODES FOR DIFFUSE POLLUTION CONTROL BY WETLANDS Chen Qingfeng1 , Shan Baoqing2 and Ma Junjian1 1 Shandong Analysis and Test Center, Jinan, 250014, China 2 Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing, 100085, China Diffuse pollution has been identified as an important cause of surface water quality degradation (Novotny, 1999). Some researches indicate that it is responsible for the transport of sediment, nutrients, heavy metals, oils, hydrocarbons and pesticides (Deletic, 1998; Schreiber et al., 2001; Lazzarotto et al., 2005). Furthermore, stormwater pollution can also have a profound effect on the ecological health of streams and reservoirs and is one of the main reasons for ecosystem degradation (Yin and Mao, 2002). There are many ecological engineering techniques, such as buffer zones, ponds, wetlands and riparian zones currently in use, and wetlands have been shown to be effective in removing pollutants from runoff water (Mitsch et al., 2002). In order to improve treatment efficiency, wetlands can be used as treatment trains. According to the flow route, the control chains can be designed in on-line and off-line treatment trains (Bardin et al., 2001; Michael and John, 2003; Shan et al., 2006; Paolo et al., 2006; Chen et al., 2007). For the on-line treatment train, all of the runoff from a storm routes through all the system structures, which are distributed on the runoff route. The treatment train may have lower pollutant removal efficiency than off-line treatment train if the system storage is not large enough to hold all the runoff from a significant storm event. For the off-line treatment train, the system structures are distributed away from the runoff route. The treatment train is designed to intercept the first flush, which has much higher concentration of pollutants in the initial runoff. The later runoff, with lower concentration of pollutants, overflows the catchment directly. The off-line treatment train requires less land area and it is an economical and effective measure for the control of runoff pollution in urban areas. Every mode plays an important role in stabilizing the adjacent ecosystems and reducing the load of runoff pollution. Email:chensdcn@163.com

Chen Qingfeng, Shan Baoqing and Ma Junjian2 In the process of diffuse pollution control, the selection of the mode is the key step. The selection of the two alternative modes for diffuse pollution control is based on concern with native topography, climate, storm water volume and available land area of the catchment (Figure 1). Figure 1. The flow chart of the two alternative modes selection for diffuse pollution control. If there is enough available land area in the catchment, then both modes can be selected. Otherwise, the offline mode may be the only choice for diffuse pollution control. Furthermore, the online mode may be the better choice if reuse of rainwater, additional biologic habitat, and aesthetics value are taken into consideration. In other conditions, the offline mode may be an effective choice for diffuse pollution control. In every mode, many ecological engineering techniques can be included. However, the application of the two modes in urban zoos has received little research attention. A detailed study was carried out from April 2003 to August 2005 in Wuhan City Zoo, which is surrounded by Moshui Lake. In this study, two catchments were selected to study the characteristics and performances of the online and offline modes in Wuhan City Zoo. For this purpose, an online pond-wetlands system in the Orangutan House Catchment, and an offline filtering ditch-pond system in the Canine House Catchment, were designed to control the small point and diffuse sources of pollution in the urban zoo. In the Orangutan House Catchment, an online pond wetlands system was used to control pollution from small point and diffuse sources. All the engineering constructions were built to adjust the flow rate of storm water and the kinetic energy of runoff on the runoff route. From upland to downstream, the landscape structures included upland grassland, orangutan house, sediment tank (ST), pond (P), the first wetland (W1) and the second wetland (W2). For the huge storage capacity of the pond-wetlands system (1071m3 ), most of runoff was able to be stored temporarily and purified by physical, chemical and biologic processes in the wetlands.

Two Alternative Modes for Diffuse Pollution Control by Wetlands 3 The online mode flow of the catchment is shown in Figure 2. Through grids, S2 was initially stored in ST on dry days. During rainfall events, all runoff, coming from S1 and S3, as well as S2, flowed through ST, P, W1 and W2 sequentially and then drained into Moshui Lake. In order to save water, the rainwater, stored in the pond-wetlands system, can be reused for flushing the animal house and irrigating the grassland. Figure 2. Online mode for diffuse pollution control in the Orangutan House Catchment. Without enough available land area for water treatment constructions, an offline filtering ditch-pond system was designed to control diffuse pollution in the Canine House Catchment. The off-line treatment train was composed of some pretreatment equipments and a filtering ditchpond system. The pretreatment equipments include a transport ditch, grids and a sediment tank. The filtering ditchpond system consists of a filtering ditch and two ponds. This system has a storage capacity of 115m3 and can store the initial 13.7mm runoff depths in a storm. Four species of hydrophytes, including Phragmites communis Trirn., Acorus calamus Linn., Alternanthera philoxeroides and Canna generalis, were planted in the ponds. The landscape structures in the catchment include upland grassland, storm transport ditch (T), Canine House, filtering ditch (FD) and ponds (P) from upland to downstream. FD was underground and rebuilt by an old flue, is 83 m in length, 0.5 m in width, and 1.2 min depth. It has three sections: sediment zone, filtration zone and storage zone. There are 9 subsections in the filtration zone and each subsection is filled with one of the following media: gravel, aluminite stone, bulky sand, cobblestone, ceramic granule, silver sand, turf, steel slag and vermiculite. All the ecological engineering constructions were finished in April 2004. According to the pollution characteristics, topography, available land area and climate in the catchment, the off-line treatment train was designed to separate the first flush from the runoff. The sketch map of the off-line treatment train is shown in Figure 3. Because the main type of land use is upland (61.4%) in the catchment, the off-line treatment train works in a natural process and requires no power. Through the grids, wastewater from flushing the animal houses (S2) was initially stored in the sediment tank (ST) and overflowed to filtering ditchpond system for decontamination on dry days. During rainy days, the initial runoff, coming from upland runoff (S1) and roof runoff (S3), as well as wastewater (S2), was diverted to the filtering ditch (FD) for filtration and adsorption. After that, the runoff water overflowed into ponds for further decontamination

Chen Qingfeng, Shan Baoqing and Ma Junjian4 and then, in the final stage, drained into Lake Moshui. The later runoff, with lower concentration of pollutants, was discharged into the lake directly. Figure 3. Offline mode for diffuse pollution control in Canine House Catchment. The results showed that the two modes both improved runoff water quality and had high retention rates for water and pollutants. In the outflows, the event mean concentrations (EMCs) of total suspended solids (TSS), chemical oxygen demand (COD), total nitrogen (TN) and total phosphorus (TP) were reduced by 88%, 59%, 46% and 71% for the online mode, and those were 75%, 50%, 50% and 74% for the offline mode. The annual retention rates of pollutant loads for the online mode were 94.9%98.5% in the three study years; those for the offline mode were 70.5%86.4%. Based on calculation, the online mode was able to store the runoff of 66.7 mm rainfall completely, and the offline mode could store that of 31.3 mm rainfall. In addition, the online mode can provide an effective way for rainwater utilization and good habitats for aquatic wildlives, and has an excellent aesthetics value for recreationsal pastimes. The offline mode can save land resources and may be an effective and economical measure for diffuse pollution control in urban areas. REFERENCES Bardin, JP, Barraud, S, Chocat, B, 2001. Uncertainty in measuring the event pollutant removal performance of on-line detention tanks with permanent outflow. Urban Water 3, 91106. Chen QF, Shan BQ, Yin CQ, Hu CX. An off-line Filtering Ditch-pond system for Diffuse Pollution Control at Wuhan City Zoo. Ecological Engineering, 2007, 30(4):373-380. Chen QF, Shan BQ, Yin CQ, Hu CX. Two Alternative Modes of Diffuse Pollution Control in an Urban Tourist Area. Journal of Environment Science, 2007, 19(10):1067-1073. Deletic, A, 1998. The first flush load of urban surface runoff. Water Res. 32 (8), 24622470. Lazzarotto, P, Prasuhn, V, Butscher, E., Crespi, C., Flu hler, H., Stamm, C., 2005. Phosphorus export dynamics from two Swiss grassland catchments. J. Hydrol. 304, 139 150. Michael Jr., John, H., 2003. Nutrients in salmon hatchery wastewater and its removal through the use of a wetland constructed to treat off-line setting pond effluent. Aquaculture 226, 213225. Mitsch, WJ, Lefeuvre, JC, Bouchard, VB, 2002. Ecological engineeringapplied to river and wetland restoration. Ecol. Eng. 10, 119130. Novotny, V, 1999. Integrating diffuse pollution control and water body restoration into watershed management. J. Am. Water Resour. Assoc. 35 (4), 717727.

Two Alternative Modes for Diffuse Pollution Control by Wetlands 5 Paolo, SC, Gaspare, V, 2006. Simulation of the operation of detention tanks. Water Res. 40 (1), 8390. Schreiber, J.D., Rebich, R.A., Cooper, C.M., 2001. Dynamics of diffuse pollution from US southern catchements. Wat. Res. 35(10), 25342542. Shan BQ, Chen QF, Yin CQ, 2006. On-line control of stormwater pollution by pond-wetlands composite system in urban tourist area[J]. Acta Scientiae Circumstantiae, 26(7): 1068 1075 (in Chinese). Yin, CQ, Mao, ZP, 2002. Nonpoint pollution control for rural areas of China with ecological engineering technologies. Chin. J. Appl. Ecol. 13 (2), 229232 (in Chinese).

In: Wetlands: Ecology, Conservation and Restoration ISBN: 978-1-60456-995-7 Editor: Raymundo E. Russo 2008 Nova Science Publishers, Inc. Short Communication MULTIANGULAR IMAGING OF WETLANDS IN NEW ENGLAND Lesley-Ann L. Dupigny-Giroux and Eden Furtak-Cole University of Vermont, Department of Geography 200 Old Mill Building, Burlington, VT 05405-0114, 802-656-2156 ABSTRACT Multiple view angles (MVA) or multiangular imaging represents a yet to be explored use of the remote sensing of wetlands. The ability to view the landscape off-nadir (traditionally the surface is viewed at right angles) allows for the quantification of moisture stress, species separation and the proportion of vegetation to standing water in these ecosystems. This commentary will focus on the ratio of two broadband wavelengths (near-infrared to blue) derived from multiangular images acquired by the Airborne Multi- angle Imaging SpectroRadiometer (AirMISR) of wetlands across New England. The resulting insights into the photointerpretation, monitoring and mapping of wetlands will be highlighted. 1. INTRODUCTION Multiple view angles (MVA) or multiangular imaging of terrestrial ecosystems has been shown to provide multispectral data not observed from the nadir or other single view angles only, due to the highly anisotropic reflectance of vegetation (Asner et al., 1998). Vegetation parameters may not be the most sensitive to the nadir view angle (Privette, 1995). Other studies have explored the relationship between a sensors field of view and vegetation structure (Widlowski et al., 2004; see Diner et al., 2005 for a full description of these studies), land cover classifications (Hyman and Barnsley, 1997) and the role of sub-pixel heterogeneity (Zhang et al. (2002a, b), as well as view angle and reflectance anisostropy at the red wavelengths (Pinty et al., 2002). E-mail: ldupigny@uvm.edu

Lesley-Ann L. Dupigny-Giroux and Eden Furtak-Cole8 In a recent study (Dupigny-Giroux, 2007), multiangular images from the Airborne Multi- angle Imaging SpectroRadiometer (AirMISR) of the Howland Forest in Maine, were used for land use/land cover (LULC) separability under varying moisture conditions in the humid, continental environment of central Maine. The study extended original work by Dupigny- Giroux and Lewis (1999) that used the ratio of near-infrared/blue wavelengths plotted against surface temperatures to describe vegetation and moisture stress for the semiarid Brazilian nordeste (northeast). This study complements work by Silva Xavier and Soares Galvo (2005) who used Principal Components Analysis of Multi-angle Imaging SpectroRadiometer (MISR) data from the Amazon to discriminate land cover types. Results of the Dupigny-Giroux (2007) indicated that the NIR/blue ratio at multiple view angles was able to discriminate variations among wetland, aquatic vegetation and the extent of moisture stress. Contributions of the study included an expansion of the recommended combination of the 15-30 solar illumination angle and nadir viewing angle for optimally recording benthic features (Dobson et al., 1995); a sensitivity of the NIR/blue ratio to species type and vigour, water/vegetation proportions and moisture gradients across emergent wetlands; and the distinction between aquatic macrophytes and terrestrial vegetation that are often similar individual wavelengths (Valta-Hulkkonen et al., 2003). The study suggested potential uses of the multi-angular ratio including improved mapping of wetlands in humid temperate regions (Bicheron et al., 1997; Barnsley et al., 1997); the avoidance of false change detection due to drought or water draw down (U.S. Fish and Wildlife Service, 2004) and; the improved photointerpretation of evergreen forested wetlands and more xeric ecosystems (Tiner, 2003). In this commentary, the methodology of the Dupigny-Giroux (2007) study was applied to wetlands at two other experimental forests in New England to explore the applicability of the technique across disparate wetland types and microclimates. 2. DATA AND METHODOLOGY 2.1. AirMISR Program TheAirMISR instrument is a pushbroom imager that is mounted on the NASA ER-2 aircraft flying at an altitude of 20km over selected temperate and tropical study areas. It used a single camera on a pivoting gimbal mount to collect data at the nine viewing angles used on the spaceborne MISR instrument. These angles are nadir (An), 26.1 fore (Af) and aft (Aa), 45.6 fore (Bf) and aft (Ba), 60.0 fore (Cf) and aft (Ca) and 70.5 fore (Df) and aft (Da). The swath width of the imagery varied from 11km at nadir to 32km for the D cameras. Four spectral bands were centered at 446nm, 558nm, 672nm and 867nm (blue, green, red and near- infrared) (Diner et al., 1998). The data used in this commentary were collected over three experimental forests in New England in August 2003 (Figure 1). The Bartlett Experimental Forest in north-central New Hampshire and the Harvard Forest in western Maine were both flown on 24 August, with data acquisition over Howland Forest in central Maine on 28 August. All three sites are well instrumented with standard meteorological equipment, biomass and carbon sequestration

Multiangular Imaging of Wetlands in New England 9 measurements to support long term experiments including NASAs Forest Ecosystem Dynamics Project. Howland Forest Bartlett Forest Harvard Forest Figure 1. Locations of the three experimental forests in New England. Only the data from the north-south runs and A-C cameras over each site were used due to data inhomogeneities. Georectified radiance product (L1B2) data were resampled to a 27.5m grid in the UTM (Universal Transverse Mercator) projection and available online from the Langely Distributed Active Archive Center (DAAC). Actual radiances were computed using the AirMISR tool. An IR minimum check was performed for each viewing angle. Ancillary digital data were acquired from the National Wetlands Inventory, Maine GIS, New Hampshire GIS and Massachusetts GIS. 2.2. Wetlands of the Study Sites The wetlands observed at the three study sites varied by extent, species composition, tidal regimes and permanence of water. The Howland Forest site decreases in elevation from over 120m in the north to about 19m in the south, with palustrine, estuarine, evergreen as well as broad-leaf deciduous and persistent emergent wetlands. To the west, the Bartlett Experimental Forest site located in the White Mountains National Forest ranges in elevation from 59m to 1868 m, an upland area characterized by broadleaf deciduous forests with predominantly palustrine forested broad-leaf deciduous and needle-leaf evergreen wetlands. The Harvard Forest site was lower in elevation (9-548m) and characterized by both palustrine evergreen and freshwater forested shrub wetlands.

Lesley-Ann L. Dupigny-Giroux and Eden Furtak-Cole10 3. RESULTS AND APPLICABILITY OF MVA TO WETLAND STUDIES The relationship between wetlands and view angles can be analyzed by scatterplots (not shown) of camera pairs on which the 45 and best fit lines have been plotted. For the three experimental forests, there was a high degree of correlation between the high camera view angles (An, Af, Aa and Bf). The straight line relationship denotes a moisture gradient from mesic regions (low ratios) to xeric ones (high ones). The relationship is most extensive (with points at both ends of the 1:1 line) for the palustrine, estuarine wetlands of the Howland forest and less so for the other two regions. Differences in species composition and tidal flow regimes were marked across the three forests, influencing the view angles that were most useful for wetland discrimination. For example, at the Howland Forest wetlands, the scatterplot of the Af and Bf forward viewing angles (R2 =0.904) was particularly well suited to highlighting moisture stress across forested wetlands, stress that was not observable at nadir (Dupigny-Giroux, 2007). This may be due to the fact that these seasonally flooded wetlands tend to be wetter for shorter durations during the growing season (Tiner, 2003). At the Harvard Forest, both the scatterplots of the high forward view angles (Af and Bf) as well as the nadir (An) and Af pair had the most significant best fit lines (R2 values of 0.946 and 0.957 respectively). The An-Af pairing was marginally better in that it deviated less from the 1:1 line than the Af-Bf pairing. For the Bartlett Forest wetlands, the regression statistics for camera pairs were quite low (R2 80% (Brooks et al., 2000) VSSF 24% (Meuleman et al., 2003) VSSF 59.5% (Vymazal, 2007) VSSF (pilot-scale) < 47.4 % (Sleytr et al., 2007) Phosphorus may also be bound to the substrate of the SSF mainly as a consequence of adsorption and precipitation reactions with calcium, aluminum and iron in the substrate. The capacity of a CWS to remove phosphorus from wastewaters may then be dependent of substrate characteristics (contents in Al, Fe and Ca ions, grain size distributions, pH and specific area). Various artificial media have been tested in order to improve the P-removal in CWS among which are, for example, light expanded clay aggregates (LECA), wollastonite, vermiculite, diatomaceous earth, blast furnace slag and limestone (Brooks et al., 2000; Johansson and Gustafsson, 2000; Brix et al., 2001; Oovel et al., 2007). The removal of phosphorus through adsorption and precipitation can be significant (Vohla et al., 2005) but it is important to realize that these processes are saturable and adsorption decreases over time (Vymazal, 2007). In addition, daily pH variations due to the respiration/photosynthesis cycles may be responsible for cycles of phosphorus precipitation/resolubilization. 4.4. Pathogen Removal Waterborne diseases remain a major hazard in many parts of the world. The important organisms from a public health point of view are the pathogenic bacteria and viruses. Protozoan pathogens and helminth worms are also of particular importance in tropical and subtropical countries (Rivera et al., 1995; Cooper et al., 1996; Vymazal et al., 1998b).

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto46 Wastewater discharges are the major source of contamination by faecal pathogenic microorganisms in rivers and coastal waters posing a risk to public health (Mason, 2002; Sleytr et al., 2007). The treatment of wastewater pathogens in CWS is essentially a two stage process. Most pathogens are particles ranging from very small viruses to the large eggs and cysts of helminths. One of the stages of pathogen treatment is particle removal. This occurs via the same processes as for removal of suspended solids, namely sedimentation, filtration, surface adhesion and aggregation. A series of other processes are important in influencing the viability of pathogens as infectious agents which may occur in a stage either before or after pathogenic particles removal. The major mechanisms in this stage are the hostility of the environmental conditions (temperature, pH, dissolved oxygen concentration, redox potential, salinity, turbity), predation by nematodes, protists and zooplankton and infection by other organisms, antibiosis, exposure to UV radiation and natural die-off (Metcalf and Eddy, 1991; Cooper et al., 1996; Kadlec and Knight, 1996; Ottova et al., 1997; Vymazal et al., 1998a). The efficiency of CWS concerning the removal of microorganisms, especially indicator microorganisms like coliforms and enterococci, is a topic that has been thoroughly investigated (Kadlec and Knight, 1996; Perkins and Hunter, 2000; Langergraber and Haberl, 2001; Hench et al., 2003). In table 8, a small sample of such studies is presented as an illustrative display of the typical efficiencies achieved by several types of CWS. Reported faecal bacteria removal efficiency in CWS is generally high, usually exceeding 85%, and is usually higher for faecal coliforms and somewhat lower for faecal streptococci (Vymazal, 2005b). Is should however be noted that, in spite of high removal efficiencies, if the number of bacteria at the inflow is very large, at the outflow bacteria number may still be too high to meet wastewater quality criteria. Treatment efficiencies depend on several design and operational parameters including the type of CWS, hydraulic regime, type of vegetation, hydraulic residence time, hydraulic and mass loading rate, substrate, and temperature. The efficiency of pathogens treatment does show some variation according to the CWS type and observed efficiencies can in most cases be ranked in the order: hybrid wetlands > SSF wetlands > FWS wetlands (Vymazal, 2005b). These differences may be related to the larger contact area among water, bacteria and substrate, which is much bigger in SSF constructed wetlands compared to FWS (Sleytr et al., 2007) therefore enhancing the process rates of the system (Langergraber and Haberl, 2001). Wetland vegetation plays a crucial role in increasing the efficiency removal of pathogen in CWS. Wetland vegetation improves the trapping efficiency for small particles like viruses by increasing the surface area of biofilms in the flow path. Many species can also release exudates having antimicrobial properties or which can enhance the development in the rhizosphere of populations of bacteria with antibiotic activity (e.g. Pseudomonas).

Wetlands: Water Living Filters? 47 Table 8. Removal of pathogens by different types of CWS Indicator microorganism s Type of CWS Removal efficiencies References Faecal coliform (FC) Hybrid systems (review) 99.4 % (Vymazal, 2005b) FWS 85 94 % (Perkins and Hunter, 2000) FWS 52 % (Cameron et al., 2003) FWS > 99 % (Garca et al., 2008) FWS (review) 85.6 % (Vymazal, 2005b) Hierarchical Mosaic of Aquatic Ecosystems (HMAE) 99.997 % (Ansola et al., 2003) Set of single-family constructed wetland (review) 88 % (Steer et al., 2002) SSF 92 % (Vymazal, 2005c) SSF 91 % (Mashauri et al., 2000) SSF 93 98 % (Karathanasis et al., 2003) SSF 99 % (Garca et al., 2008) SSF 99.999 % (Soto et al., 1999) SSF (review) 91.5 % (Vymazal, 2005b) SSF > 99 % (Hench et al., 2003) Faecal streptococci (FS) Hybrid systems (review) 97.7 % (Vymazal, 2005b) FWS 82 90 % (Perkins and Hunter, 2000) FWS 99 % (Garca et al., 2008) FWS (review) 84 % (Vymazal, 2005b) SSF 99 % (Garca et al., 2008) SSF > 98 % (Mantovi et al., 2003) SSF 93 - 98 % (Karathanasis et al., 2003) SSF 83 90 % (Perkins and Hunter, 2000) SSF (review) 92.6 % (Vymazal, 2005b) SSF 99.999 % (Soto et al., 1999)

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto48 For example, it has been shown that root excretions of species such as Scirpus lacustris and Phragmites australis can diminish the populations of faecal indicators and pathogenic bacteria (Vymazal, 2005b). In addition, the presence of oxygen in the water column (produced by photosynthetic activity of submerged plants and algae in FWS, or released in the rhizosphere through the roots of macrophytes in SSF) creates unfavorable life conditions for enteric bacteria which are either facultative or obligate anaerobic. Effects of the hydraulic retention time are very simple: the longer the wastewater remains in the system, the longer bacteria remains exposed to unfavorable conditions. 4.5. Metals Removal Metals are naturally present in the environment. However, human activities are responsible for a significant increase in their concentration levels up to a point where they begin to pose an environmental and public health problem. Beyond the natural sources, contamination with metals is mainly associated with such activities as soil disturbance, mining, manufacturing, urbanization, burning of fossil fuels and use of manufactured products such as paints, pesticides, sacrificial anodes and anti-foulants. In small doses some metals are, in fact, essential to some biological processes (e.g. copper, chromium, nickel, zinc). For example, at low concentrations, copper is a micronutrient of plants essential to the photosynthetic electron transport system. However, at higher concentrations, it is marketed as an effective herbicide. In addition to concentration, the chemical form is also associated with a greater metal toxicity. For example, the methylated form of mercury is much more toxic than inorganic mercury (Mitra, 1986). Toxic effects by metals are varied and sometimes diffuse and difficult to characterize. In some cases, for very high metal concentrations, toxicity may be acute and ultimately lethal. However, usually toxicity by metals will cause chronic effects resulting from a long-term exposure. Examples of chronic health effects include cancer, disruption of the endocrine system, liver and kidney damage, disorders of the nervous system, damage to the immune systems, and birth defects. Some metals are not easily eliminated by the organisms and, therefore, they have the potential for bioaccumulation and biomagnification. This constitutes one of the major problems with metal contamination, with the potentiation of the metals' chronic toxicity along the food chain. In CWS a variety of processes may provide routes for metal retention in the CWS components and their elimination from the wastewater. The main mechanisms occurring in each of the compartments (solid medium, aqueous medium and vegetation) are illustrated in figure 4. The substrate is generally considered to be a sink for metals anthropogenically introduced in the environment. A major fraction of the elements entering the CWS will rapidly be adsorbed onto the solid phase, where a number of physical and chemical processes will determine the strength of metal retention. A small proportion of the metals can, however, remain dissolved and become available for plant uptake. In CWS another important role in metal removal is played by plants through several processes which include filtration, adsorption, cation exchange, uptake, and root-induced chemical changes in the rhizosphere (Dunbabin and Bowmer, 1992; Chen et al., 2000; Vandecasteele et al., 2005).

Wetlands: Water Living Filters? 49 Figure 4. Metal removal mechanisms in CWS (adapted from Cooper et al. (1996)). Long-term metal sequestration by plants depends on the rate of uptake, rates of translocation and retention within individual tissue types, and the rate and mode of tissue decomposition (Kadlec and Knight, 1996). Studies report the highest amounts of metals in the roots, while leaf tissue has the second highest concentrations followed by stems and rhizomes (Burke et al., 2000). Microorganisms may also play a relevant role in heavy metal removal. Such contribution may occur through their metabolism with the modification of the oxidation states of metals which in turn may lead to other transformations, such as precipitation, that effectively remove them from the wastewater. A more detailed account follows of the several physical, chemical and biological processes which concur for heavy metal removal in a CWS. 4.5.1. Physical Removal Processes Sedimentation this has long been recognized as one of the main processes in removal of heavy metals from wastewater in natural and constructed wetlands (Kadlec and Knight, 1996; Hammer, 1997; ITRC, 2003). Sedimentation is a physical process which follows other mechanisms (precipitation/co-precipitation and flocculation) whereby heavy metals aggregate into particles large enough to sink (Walker and Hurl, 2002). In this way heavy metals are removed from wastewater and trapped in the wetland sediments, thus protecting the ultimate receiving water bodies, i.e. aquatic ecosystem (Sheoran and Sheoran, 2006). Efficiency of sedimentation is proportional to the particle settling velocity and the length of wetland. 4.5.2. Chemical Removal Processes In addition to physical removal processes a wide range of chemical processes are involved in the removal of heavy metals in the wetlands:

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto50 Sorption among the chemical processes, sorption is one of the most important removal processes in wetlands, which results in the transfer of ions from water to the soil/sediments and a short-term retention or long-term immobilization of several classes of contaminants (Sheoran and Sheoran, 2006). In sediments heavy metals are adsorbed by either cation exchange or chemisorption (Sheoran and Sheoran, 2006). In the former process the metal cation will exchange with other small cations (such as, Na+ , K+ , NH4 + , etc) in their positions in the mineral structures of clays and negatively charged groups of humic acids. Therefore the capacity of soils for retention of metal cations, expressed as cation exchange capacity (CEC) increases with increasing content in clay and organic matter. Chemisorption is a process which involves the formation of chemical bonds with the surface, frequently through complexation/chelation phenomena. The adsorption capacity by cation exchange or non-specific adsorption depends upon the physico-chemical environment of the medium (e.g. pH, the properties of the metals concerned and the concentration and properties of other metals and soluble ligands present) (Debusk et al., 1996; Sheoran and Sheoran, 2006). Therefore, heavy metals speciation may change with time as the sediment conditions change (Groudev et al., 1999; Wiebner et al., 2005; Sheoran and Sheoran, 2006). Much of the heavy metals can be easily adsorbed onto particulate matter in the wetland and subsequently be removed from the water by sedimentation. Lead and copper in general tends to be adsorbed most strongly while zinc, nickel and cadmium are usually held weakly which implies that these metals are likely to be more labile and bio-available (Sheoran and Sheoran, 2006).The adsorption of metals varies with the fluctuation of pH in the outflow water (Machemer and Wildeman, 1992). Precipitated hydroxides may also act as adsorption sites for phytotoxic metals present in the water compartment of the wetland (Wood, 1990). Oxidation and hydrolysis of metals The states of oxidation of a metal will have a marked influence in its chemical behavior in water. In particular, under some oxidation states a metal may typically hydrolyze to form insoluble oxides or hydroxides whereas in other oxidation states it can be more soluble. Such is the case, for example, of iron, aluminum and manganese which can form insoluble compounds through hydrolysis (sometimes following oxidation processes). This leads to the formation of a variety of oxides, hydroxides and oxyhydroxides (Woulds and Ngwenya, 2004; Sheoran and Sheoran, 2006) Iron removal depends on pH, redox potential and the presence of various anions. In alkaline conditions Fe2+ is a highly soluble cation in water with low content of dissolved oxygen. On the other hand, the form Fe3+ is insoluble except in very acid conditions (pH < 3.5). Manganese removal is the most difficult to be achieved because its oxidation takes place at a pH close to 8 (Stumm and Morgan, 1981). In this case bacteria play an important role in the oxidation of Mn by catalyzing the oxidation of Mn2+ to Mn4+ . On the other hand, aluminum removal is purely governed by pH. Aluminum hydroxides will precipitate at pH above 5.0-6.0. Precipitation and co-precipitation is a major process of heavy metals removal in wetland sediments. The formation of insoluble metal precipitates is one of many factors limiting the bioavailability of heavy metals to many aquatic ecosystems. Precipitation depends on the solubility product (Ks) of the metal involved, pH of the wetland and concentration of metal ions and relevant anions (Brady and Weil, 2002).

Wetlands: Water Living Filters? 51 Co-precipitation is an adsorptive phenomenon also frequent in wetland sediments. Heavy metals commonly co-precipitate with secondary minerals. Copper, nickel, manganese, and other metals are co-precipitated in Fe oxides and cobalt, iron and nickel are co-precipitated in manganese oxides (Stumm and Morgan, 1981). In addition arsenic and zinc were reported to be retained on iron plaques at the surface of plant roots (Otte et al., 1995). Oxiferric hydroxide surfaces are positively charged under acidic pH conditions and negatively charged under alkaline pH conditions. Thus, adsorption and removal of oxyanions such as arsenic, antimony and selenium, through iron co-precipitation, is favored under acidic pH conditions (Brix, 1993). Alkaline conditions are favorable for co-precipitation of cationic metals such as copper, zinc, nickel and cadmium. Thus metals may become associated with iron and manganese oxides as a result of co-precipitation and adsorption phenomena (Stumm and Morgan, 1981). The process is presumed not to be very important in long-term removal and retention of metals because iron and manganese oxides, being redox sensitive, many re- dissolve following changes in oxygen concentration (Sheoran and Sheoran, 2006). In addition to oxides, hydroxides and oxihydroxides resulting from hydrolysis, other typical insoluble metal compounds include carbonates and sulfides. Conditions exist for precipitation of heavy metal carbonates when the bicarbonate concentration in water is high. Carbonate formation can take place when bacterial production of bicarbonate alkalinity in wetland sediments is substantial (ITRC, 2003). Carbonate precipitation is especially effective for the removal of lead and nickel (Lin, 1995), but Sobolewski (1999) some authors reports significant quantities of copper and manganese carbonates accumulated in some natural wetlands. Wetlands with appropriate substrate may promote the growth of sulfate reducing bacteria in anaerobic conditions. These bacteria will generate hydrogen sulfide which reacts with most heavy metals leading to formation of highly insoluble metal sulfides (Stumm and Morgan, 1981). These provide for long-term metal removal, remaining permanently in wetland sediments as long as they are not re-oxidized (Sobolewski, 1999). Metals such as copper, lead, zinc, cadmium, and arsenic may form highly insoluble sulfides in contact with low concentration of H2S (ITRC, 2003). Field results suggest that upon start up of a constructed wetland, the adsorption of dissolved metals onto organic sites in the substrate material will be an important process but over time sulfide precipitation becomes the dominant process for metal removal (Machemer and Wildeman, 1992). 4.5.3. Biological Removal Processes Biological removal is perhaps the most important pathway for heavy metal removal in the wetlands where plant uptake plays probably the most widely recognized role. While sediments of wetlands form primary sinks for heavy metals (Gray et al., 2000), macrophytes may absorb heavy metals through roots and also shoots. It has been proposed that the processes used by plants are not necessarily the same for different species and for different metals. Submerged rooted plants may have high potential for the metals phytoextraction from sediments as well as water, while floating plants can extract metals only from water (Sriyaraj and Shutes, 2001). Among such processes can be mentioned: sorption by roots (a combination of physical and chemical processes such as chelation, ion exchange and chemical precipitation), and the biological processes including

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto52 translocation to the aerial part and precipitation induced by root exudates or by microorganisms. The rate of metal removal by plants varies widely, depending on plant growth rate and concentration of the heavy metals in plant tissue. The rate of metal uptake per unit area of the wetland is often much higher for herbaceous plants, or macrophytes such as duckweed (Lemna minor) (Zayed et al., 1998), salix (Stoltz and Greger, 2002), cattail (Typha latifolia) and common reed (Phragmites australis) (Sheoran and Sheoran, 2006). Some of these species can tolerate high concentrations of several metals in their body mass without showing negative effects on the growth (Sheoran and Sheoran, 2006). There are also some examples in the literature indicating that some species may have the ability to accumulate only specific heavy metals, e.g. the Spirodela polyrhiza for Zn (Markert, 1993). Constructed wetlands with well grown Cyperus alternifolius and Vallarsia exaltata have been reported to be an effective tool in phytoremediation of cadmium, copper, manganese, zinc and lead (Cheng et al., 2002). Microorganisms also provide a measurable amount of heavy metal uptake and storage; it is their metabolic processes that play the most significant role in removal of heavy metals (Ledin and Pedersen, 1996; Russell et al., 2003; Hallberg and Johnson, 2005; Sheoran and Sheoran, 2006). Reduction of metals to non-mobile forms by microbial activity in wetlands has been reported by Sobolewski (1999). Metals like chromium and uranium become immobilized when reduced through processes biologically catalyzed by microorganisms (Fude et al., 1994). In table 9 are presented illustrative results found in the literature concerning the removal efficiencies for several heavy metals obtained in different constructed wetlands systems. Table 9. Heavy metals removal efficiencies in different types of CWS Pollutant Type of CWS Reduction rate References CWS (small-scale plot) 90% (Liu et al., 2007)Lead HSSF 70% (Mantovi et al., 2003) FWS 67 % (Maine et al., 2006) FWS 48 % (Maine et al., 2007) Grenhouse experiment > 43 % (Hadad et al., 2007) HSSF 59 % (Mantovi et al., 2003) HSSF 49 % (Lesage et al., 2007) Nickel VSSF 80% (Lee and Scholz, 2007) HSSF 79 % (Mantovi et al., 2003)Copper VSSF 95 % (Lee and Scholz, 2007) Iron FWS 95 % (Maine et al., 2006) Aluminum HSSF 93% (Lesage et al., 2007)

Wetlands: Water Living Filters? 53 Pollutant Type of CWS Reduction rate References FWS 58 % (Maine et al., 2007) FWS 86 % (Maine et al., 2006) Grenhouse experiment 100% (Hadad et al., 2007) Chromium HSSF 52% (Mantovi et al., 2003) CWS (small-scale plot) 90% (Liu et al., 2007) Grenhouse experiment 35% (Hadad et al., 2007) Zinc HSSF 86 % (Mantovi et al., 2003) CWS (small-scale plot) 90% (Liu et al., 2007)Cadmium HSSF 24 % (Mantovi et al., 2003) Removal efficiencies reported in CWS studies present some variation, from quite low values (~ 25%) to nearly complete removal of some metals. In general, however, the efficiencies are high (> 70%) but these will depend, as usual, on varied factors such as the influent metal loads, the type of vegetation used, the CWS type and on environmental conditions. Obviously, better removals will be achieved when the systems are specifically designed and optimized to solve well-defined metal contamination problems such as mine drainage, where well-known metal accumulator plants will be used preferably, in comparison with systems designed for broader treatment targets where metals are only possibly one among several types of pollutants to remove from wastewaters. As elemental substances, metal cations are naturally non-biodegradable and, for their permanent elimination from the system, the portion of metals removed by plant uptake will require a periodic plant harvesting. For CWS designed to treat high loads of metal inputs, the harvested plant biomass should afterwards be disposed as hazardous waste and receive appropriate treatment. 4.6. Organic Xenobiotics Removal Organic xenobiotics include a large range of synthetic organic compounds, such as phthalates, polychlorinated biphenyls (PCBs), dioxins, polycyclic aromatic hydrocarbons (PAHs), pesticides, sulfonated azo dyes, alkylphenols, bisphenols and pharmaceuticals and personal care products (PPCPs) (Wu, 1999; Mason, 2002; Fent et al., 2006). Several of these substances have been released in increasing amounts in the environment since decades, and, due to the low degradation rate of many of these compounds, a significant increase of their background concentrations has been observed in the different environmental compartments (Tyler et al., 1998; Skakkebaeck et al., 2000). A growing environmental concern has been emerging in recent years, because of the high toxicity and high persistence of most of these substances in the environment and in biological systems. Even though they occur only at very low concentrations in the environment, and their threats to aquatic life and public health are still not completely understood, nevertheless, sub-lethal effects of these compounds over long-term exposure may cause significant damage to aquatic life, particularly considering that some of these compounds may cause significant

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto54 endocrine disruption, impair reproduction functions of animals or even be carcinogenic, mutagenic or teratogenic (Wu, 1999; Mason, 2002). Furthermore, the high lipophilicity of many of these xenobiotics greatly enhances their biomagnification, thereby posing potential health hazards on predators at higher trophic levels (including human beings). The major ecological concern of xenobiotics is their ability to impair reproductive functions and subsequently threaten survival of the species. In fact, there is growing evidence from laboratory and field studies showing that exposure to trace amounts (g/L ng/L level) of certain xenobiotic organic compounds (e.g. halogenated hydrocarbons, PCBs, DDT, TBT) may cause reduced gonad development, disruption of normal metabolism of sex hormones (including gonadotropins), arrest of sperm maturation and block a variety of oestrogen-like effects on female reproductive systems in fish, birds, reptiles and mammals. This in turn, may lead to reproductive dysfunction such as delayed sex maturity, reduced fertility and hatch rate, depression in secondary sexual characteristics, alternation of sex behavior and viability of offspring (Wu, 1999). Due to long environmental and biological half lives, recovery from the effects of many xenobiotic compounds is expected to be slow. Indeed, it has been shown that some 15 years were required to remove the negative effects of DDT on reproduction of the white tail eagles in the Baltics, and another 10 years for the population to recover (HELCOM, 1996). Furthermore, despite a decrease in environmental concentrations, the adverse effects may remain in the ecosystem for a much longer period. In the Baltics, DDE decline to 10% of the original levels in 1984, but increased again afterwards, and the egg shells of fish eating birds, which had begun to return to normal, have recently become thinner again. Thus, the downward trend was halted after the ban, and may be due to the recycling of persistent chemicals in sediment (HELCOM, 1996). A variety of sources may be the origin for the presence of organic xenobiotics in water bodies. A number of xenobiotics classes (phthalates, pesticides, PCBs and bisphenols) are industrial products, used worldwide in several applications and are therefore ubiquitous pollutants (Safe, 1994; Stales et al., 1997; Mason, 2002). Other kinds of compounds (dioxins and PAHs) are not commercial products, but are formed as by-products of various industrial and combustion processes; they are transported from atmosphere to soil and water bodies by the atmospheric runoff or deposited on the soil during the dry period and then go through the water cycle by land runoff (Birkett and Lester, 2003). Several studies have shown that a vast range of these xenobiotics are present in the effluents from domestic and industrial conventional wastewater treatment plants (WWTPs) (Birkett and Lester, 2003) which indicates that they resist removal by conventional wastewater treatment processes and may persist in the environment even after going through WWTPs. These are designed to deal with bulk substances that arrive regularly in large quantities (TSS, organic matter and nutrients) and many of these organic xenobiotics show a different chemical behavior for which the conventional processes are not well-suited. The different studies show that the WWTPs removal rates vary according to compound nature, WWTPs overall performance, and environmental conditions. This is also of potential concern about treated wastewater reuse for non industrial applications, such as irrigation of crops and aquaculture, since these pollutants may become a source of contamination of the food chain.

Wetlands: Water Living Filters? 55 4.6.1. Removal Processes in CWS Organic xenobiotics removal by CWS involves several interdependent processes which may be classified as abiotic (physical or chemical) or biotic (microbial or phytological). The primary abiotic and biotic processes that participate in removing organic xenobiotics from contaminated water in a CWS are described in table 10 (Evans and Furlong, 2003; Pilon-Smits, 2005). Table 10. Abiotic and biotic processes involved in xenobiotics removal in CWS Processes Description Abiotic Sorption Including adsorption and absorption, the chemical processes occurring at the surfaces of plants and substrate that result in a short-term retention or long-term immobilization of xenobiotics Hydrolysis The chemical breakdown of organics by the action of water, a process which frequently is pH-dependent Photodegradation/ oxidation Degradation/oxidation of organic xenobiotics by the action of sunlight Oxidation/reduction Modification, which sometimes may be quite substantial, of the xenobiotics due to the action of oxidizing (frequently dissolved oxygen) or reducing agents. Sometimes a redox reaction is a first step leading to removal by other processes, such as precipitation or volatilization. Redox reactions are also frequently brought about by biotic agents such as bacteria, or enzymatically catalyzed Precipitation Many organic compounds have low water solubility and, especially those exhibiting acid-base properties, may convert into insoluble forms by pH changes Settling and sedimentation Removal of particulate matter and suspended solids Volatilization Release of some organic xenobiotics, as vapors, which occurs when these compounds have significant vapor pressures Biotic Aerobic/anaerobic biodegradation Metabolic processes of microorganisms, which play a significant role in organic xenobiotics removal in CWS Phytodegradation Breakdown of organic xenobiotic, either internally, having first been taken up by the plants, or externally, using enzymes excreted by them Rhizodegradation Plants provide root exudates that enhance microbial degradation of some organic xenobiotics Phytovolatilization/ evapotranspiration Uptake and transpiration of volatile organic xenobiotics through the leaves The contribution of each process to the overall efficiency of the system will be very dependent on a wide variety of factors relative not only to each CWS component characteristics but also to the properties of the organic xenobiotics, the characteristics of the wastewater and the environmental conditions.

Ana Dordio, A. J. Palace Carvalho and Ana Paula Pinto56 4.6.1.1. Factors Affecting Organic Xenobiotics Removal Efficiency in a CWS The degree to which each process will contribute to the overall removal of the organic xenobiotics from contaminated waters in CWS is in turn dependent on the physico-chemical properties of these compounds (e.g., water solubility, sorptive affinity), characteristics of the substrate (e.g., pH, organic matter content, redox status), the plants species, effluent characteristics (e.g., pH, dissolved organic matter, electrolyte composition) as well as other environmental conditions (e.g., temperature, moisture). Some of the most important organic xenobiotics properties that affect their behavior and removal in CWS are its molecular structure, polarity, ionization constant (pKa), water solubility, sorption coefficient (Kd), octanol-water partition coefficient (Kow), volatility and chemical stability. The texture of a soil or substrate is an extremely important characteristic in the sorption process. If the substrate is made up of mostly clay and organic matter a significant amount of sorption will take place. Clay, especially intermixed with organic particles, by far adsorbs the most out of the main types of texture (e.g. silt and sand) because of its small particle size, high surface area and high surface charge. The content in organic matter also has a strong influence in the sorptive properties of the mineral media, mainly due to the presence of humic acids which form a large portion of their composition (Brady and Weil, 2002). These huge organic molecules are characterized by hydrophobic regions suitable for the

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