abstracts oral presentation posters · comparing outflows of the individual treatments for 2013,...

ABSTRACTS - WETPOL 2013 - October 13-17, 2013 - Nantes - FRANCE 3

Abstracts Oral presentation Posters

The first part of the abstracts section is dedicated to

Keynote speakers and Invited speakers.

Submissions are identified both by a number and an

abbreviation as mentioned on the programme : Keynote

Speaker “K”, Invited Speakers “IS”, Oral presentation

“O”, Poster and oral presentation “PO” and

Posters“P”.


KEYNOTE SPEAKERS

DR WILLIAM J. MITSCH

Protecting the Florida Everglades Wetlands with Wetlands - Can

stormwater phosphorus be reduced to oligotrophic conditions ?

Dr. William J. Mitsch is Eminent Scholar and Director, Everglades

Wetland Research Park, and Sproul Chair for Southwest Florida Habitat

Restoration and Management at Florida Gulf Coast University. He is

also Professor Emeritus of Environment and Natural Resources, The

Ohio State University, where he taught for 27 years. In August 2004 he

was awarded, along with his Denmark friend Sven Erik Jørgensen, the

2004 Stockholm Water Prize by King Carl XVI Gustaf of Sweden for

lifetime achievements in the modeling, management, and conservation of lakes and wetlands.

Dr. Mitsch holds an M.E. and Ph.D. in environmental engineering science and systems

ecology at University of Florida. His research and teaching have focused on wetland ecology

and biogeochemistry, wetland creation and restoration, ecological engineering and ecosystem

restoration, and ecosystem modeling. He has authored or co-authored over 600 publications,

reports, and books, including 4 editions of the popular textbook Wetlands. He is editor-in-

chief of the international journal Ecological Engineering and was Chair of the 1992

INTECOL Wetland Conference and EcoSummit 2012, both held in Columbus USA. Dr.

Mitsch’s other awards include two Fulbright Fellowships, the U.S. EPA National Award for

Wetland Research (1996), a Fellow of the American Association for the Advancement of

Science (AAAS) (1997), Theodore M. Sperry Career Award from the Society of Ecological

Restoration International (2005), the Lifetime Achievement Award from the Society of

Wetland Scientists (2007), and an Einstein Professorship from the Chinese Academy of

Sciences (2010).

http://fgcu.edu/swamp

PR JACQUES BRISSON

Ecoystem services of wetlands : does plant diversity really matter ?

Jacques Brisson is professor of plant ecology at the Institut de

recherche en biologie végétale of the University of Montreal. Over the

last 15 years, his research interests have focused on plant invasion,

ecosystem restauration and on the role of plants in treatment wetlands.

He has authored more than 60 articles in peer-reviewed journals, and

he is currently associate editor of the journals Botany and Ecoscience.

He is also highly involved in public education, and has authored more

than 100 articles in popular science journals. He is president of the

Quebec Society of Phytotechnology, which he founded in 2008.

Prof. Jacques Brisson

Institut de recherche en biologie végétale

Université de Montréal

4101 est, rue Sherbrooke, Montreal (Qc) H1X 2B2

[email protected]

www.irbv.umontreal.ca/personnel/chercheurs/jacques-brisson?lang=en


DR. LARS DUESTER

Wastewater, examples on new organic contaminants, upcoming metal(loid)s, nano

materials & the transfer/transformation in wetlands

Lars Duester was born 1973 in Germany. After studies in

environmental sciences and a PhD in Chemistry (2007, University of

Duisburg-Essen, Germany) he worked as a Post Doc in the Working

Group, Soil Geography and Soil Sciences at the University of

Cologne and at the Institute of Environmental- and Soil-Chemistry at

the University of Koblenz-Landau. At the end of 2010 he joined the

Federal Institute of Hydrology, research group: Aquatic Chemistry, in

Germany. His current research interests are the fate of metal(loid)s in

surface waters and associate biotopes - either released on natural basis

or from anthropogenic causes (e. g., construction materials in

hydraulic engineering). Main areas of interest are the border zone between particles/colloids

to dissolved chemical species, the transformation of inorganic into organic metal(loid) species

as well as changes in speciation and availability across aquatic interfaces.

http://www.bafg.de/cln_033/nn_929770/EN/Home/homepage__en__node.html?__nnn=true

PR. JOAN GARCIA

The Cartridge Theory: a Conceptual Approach to Horizontal-Flow Wetlands’

Functioning

Joan Garcia is Full Professor of Environmental Engineering and Director of

the Department of Hydraulic, Maritime and Environmental Engineering of

the Universitat Politècnica de Catalunya (UPC). Joan obtained his degree in

Biology in 1990 and presented his doctoral dissertation on wastewater

treatment engineering in 1996. For his contributions to water resources

research and wastewater engineering received in 2008 an award of the

Spanish Ministry of Science for the Intensification of Research Activity. He

has written over 150 articles in scientific journals.

J. García carries out interdisciplinary research on ecoinnovative treatment

systems – a new generation of sustainable environmental technologies that mimicking nature

and maximizing eco-efficiency allow treatment of wastewaters and other kinds of wastes. He

has worked in the development of ecotechnologies for wastewater treatment and at the same

time materials and energy recovery, like constructed wetlands and high rate algal ponds.

DR. KELA WEBER

The role and characterization of microbial communities in wetlands for water pollution

control

Dr. Kela Weber studied at the University of Waterloo where he

completed a BASc. in Environmental Engineering, as well as a

MASc. and PhD studying the temporal and spatial dynamics of

microbial communities in constructed wetlands. Dr. Weber went on

to complete postdoctoral fellowships at the Centre for Control of

Emerging Contaminants, and in the Research for Subsurface

Transport and Remediation (RESTORE) group at the University of

Western Ontario. Dr. Weber’s industry experience has been gained


through positions at XCG Consulting Ltd., Barrday Inc., Imperial Oil Ltd., and Shell Canada

Ltd.

Kela currently manages the Environmental and Bioprocess Engineering Laboratory (EBEL)

as an Assistant Professor in the Department of Chemistry and Chemical Engineering at the

Royal Military College of Canada. Kela is also a practising Professional Engineer as Vice-

President of Elementary Water Solutions Inc., a company that designs and installs water

treatment facilities for Industrial applications. http://www.weberwetlandlab.ca/


Protecting the Florida Everglades wetlands with wetlands - Can

stormwater phosphorus be reduced to oligotrophic conditions?

William J. Mitsch, Ph.D.

Eminent Scholar and Director, Everglades Wetland Research Park

Juliet C. Sproul Chair for Southwest Florida Habitat Restoration and Management

Florida Gulf Coast University, Kapnick Center, Naples, Florida USA ([email protected])

Professor Emeritus, The Ohio State University

Courtesy Professor of Soil and Water Science, University of Florida

Editor-in-Chief, Ecological Engineering

INTRODUCTION

The Florida Everglades, one of the largest and most unique wetland systems in the world,

and especially its “river of grass,” are being threatened by high-nutrient stormwater coming

from the highly fertilized Everglades Agricultural Area to the north. The main nutrient

problem is phosphorus, which causes the highly oligotrophic sawgrass (Cladium jamaicense)

in the northern Everglades to become transferred to a partially eutrophic cattail (Typha

latifolia/T. domingensis) community. Current government directives are requiring that the

total phosphorus concentration of storm water drainage be limited to 10 ppb (µg-L), the

approximate concentration of phosphorus in rainfall. While over 16,000 ha of so-called

stormwater treatement areas (STAs) or treatment wetlands have been restored to former

farmland to treat the stormwater and they are generally effective in removing 60 to 80% of the

total phosphorus, reaching the mandated 10 ppb threshold has not been achieved.

METHODS

A three-year mesocosm-scale experiment involving introducing low-nutrient effluent from

one of the STAs to mesocosm wetlands planed with Everglades-native wetland plants has

been conducted in the Florida Everglades from March 2010 through March 2013. Eighteen

flow-through mesocosms (6 m x 1 m x 1 m) constructed at the STA-1W research site near

West Palm Beach Florida. The inflow water used in this experiment comes from the outflow

area of STA-1W before passing in parallel through the eighteen mesocosms. The hydrological

loading rate (HLR) in all mesocosms was held at about 2.6 cm/day with 40 cm of water depth.

The eighteen mesocosms were randomly assigned with six different plant communities with

three replicates of each treatment, consisting of sawgrass (Cladium jamaicense); waterlily

(Nymphaea odorata); cattail (Typha domengensis); submerged aquatic vegetation (SAV)

including Najas guadalupensis, Chara sp. and a water lily-Eleocharis sp. mixed community;

and a soil without vegetation as a control. Through the first two years of the study, the control

system became dominated by Najas guadalupensis, but later it was dominated by Chara sp.

Water quality samples were collected at the main inflow from the canal and outflows of the

mesocosoms twice per month since August 26, 2010, four months after vegetation was

introduced. Nutrients, including total phosphorus (TP), dissolved inorganic phosphorus (DIP),

dissolved organic phosphorus (DOP), particulate phosphorus (PP), dissolved organic carbon

(DOC), total dissolved carbon (TOC), total dissolved Kjeldahl nitrogen (TDKN), total

Kjeldahl nitrogen (TKN), dissolved calcium (Ca2+

) and dissolved magnesium (Mg2+

) were

analyzed at the SFWMD laboratories using standard methods.

RESULTS AND DISCUSSION

Total phosphorus (TP) concentrations in the inflow water ranged between 13 and 78 µg-

P/L (average 25±1 µg-P/L, n = 55) from August 2010 through March 2013. The outflow TP


of the water lily-Eleocharis sp. mixed community treatment was much higher than the

outflows of the other treatments for the first two years (August 2010 through July 2012) and

averaged 103±12 µg-P/L over the entire study period. The outflows of the other treatments

averaged between 31±2 µg-P/L (water lily treatment) and 47±3 µg-P/L (sawgrass treatment)

for the entire study period.

The outflow TP of all of the treatments began to show decreases from the middle of

November 2011 (Figure 1). Through 2012 the average outflow of all of the treatments was

34±1 µg-P/L, a 51% decrease from the average outflow of 69±6 µg-P/L for 2011. Outflow

total phosphorus of the vegetation treatments began to converge for all treatments, including

the mixed Eleocharis/water lily treatment, in late 2012 and outflows began to be routinely

lower than the inflow about then. From November 2012 through the end of this reporting

period of early March 2013 (9 sampling periods), Seventy percent of the treatment outflow

concentrations were lower than the inflow concentrations. The average TP concentration

decreased overall to 19±1 (n =5) for 2013 a decrease of 44% from the 2012 average. This

suggests that the suspected phosphorus reflux from the mesocosm soils into the water column

slowed after 2 years of mesocosm operation and that the soil phosphorus concentrations may

have reached equilibrium with phosphorus concentrations in the water column.

Comparing outflows of the individual treatments for 2013, the only reasonable time to

compare inflows and outflows without phosphorus efflux from the soils, the water lily

treatment was lower (p<0.05) than the inflow and all of the other treatments with an average

outflow concentration of 11±1 µg-P/L. The outflow P concentrations of the control/Chara and

cattail (Typha) treatments were also significantly lower (p<0.05) than the inflow during this

time period with concentrations of 15±3 and 16±1 µg-P/L respectively. The sawgrass, SAV,

and mixed community outflow total phosphorus was not different than the inflow during the

2013 sampling.

Phosphorus was exported from the mesocosms, i.e., outflows were higher than inflows, for

2010, 2011, and most of 2012. When the 2013 data are isolated and evaluated, 4 of the 6

vegetation treatments showed total phosphorus removal, ranging from minimum to maximum

removal as mixed community (17% removal), cattail (28% removal, control (34% removal)

and water lily (51%) removal. The SAV and sawgrass treatments exported phosphorus, both

with 14% export. Overall there is a significant different between the inflow and outflow for

all the treatments (F=7.818, p=0.000) in 2013 for percent total phosphorus removed. Overall

the water lily treatment is not statistically different from the control or cattail treatment in this

small data set.

CONCLUSIONS

1. Our mesocosms are beginning to show phosphorus retention after 2 years of

phosphorus export, hypothesized to be due primarily to a reduction from efflux from

soils and possibly due to changes in water column productivity and macrophyte

biomass production over that time.

2. Evaluating the 2013 preliminary data only, phosphorus retention is most effective in

the water lily treatment, slightly less effective in the control (or “self-design”), and

cattail treatments. The other treatments (sawgrass, lily plus Eleocharis, and SAVs) do

not show statistical differences between inflows and outflows for 2013. All of these

results are preliminary because of the few data reported here for our study period in

2013 to date.

3. Our mesocosms, within the last 2 months of data, shown treatments resulting in

concentrations of total phosphorus of less than 10 ppb for some sampling periods.

4. If the recent results from the water lily treatment retention are used as an indicator, the

phosphorus retention per unit area of these additional low-P treatment wetlands is one-


tenth (0.12 g-P m-2

yr-1

) the rates of the current STAs (1.2 g-P m-2

yr-1

) at these low

inflow phosphorus concentrations. In other words, any new STAs to achieve 10 ppb

consistently would have to be designed to 10 times larger per kg of phosphorus than

the original STAs.

5. Out study concludes that any treatment wetland constructed as STAs with local soils

to achieve low (~10-15 ppb P) concentrations would probably take a minimum of 2

years to be effective. Before that time, the wetlands would probably be net sources of

phosphorus.

6. Because of the two years that these mesocosms took to deplete their labile phosphorus

in their soils, the results presented here are preliminary and need verification with a

one or two-year continuation of this mesocosm study. Without this verification, these

results remain preliminary and insufficient for any additional wetland designs.


Ecosystem services of wetlands : does plant diversity really

matter ?

Jacques Brisson

Institut de recherche en biologie végétale, Département de sciences biologiques, Université de

Montréal, 4101 est, rue Sherbrooke, Montréal (Qc) H1X 2B2, CANADA.

[email protected]

Remarkable progress has been made towards understanding how the loss of biodiversity

affects ecological functioning and services. For example, there has been accumulating

evidence that loss of biodiversity results in a reduction of the efficiency with which

communities capture resources, produce biomass and recycle nutrients. Where ecosystem

properties and functions could be related to ecosystem services such as storing and cycling of

nutrients, clear positive effects of biodiversity have been documented. However, it is not clear

how these patterns apply to all ecosystem types. There is an undisputed over-representation of

grasslands and primary production measures in the literature on this subject. Despite the

universally recognized efficiency of wetlands in cleaning water, our understanding of the

fundamental relation between plant diversity and ecological attributes and function is still

limited for this system. For example, in the most recent meta-analysis examining how species

richness influences ecological processes (Cardinale, 2011), it was shown that standing

biomass was highly correlated to the initial number of species in several ecosystem types, but

not wetlands, although the small number of experiments (7 out of 477) signals a need for

caution in interpreting these results. Similarly, lower nutrient concentration in soil or water

was correlated with diversity, but supporting evidence was essentially based on nutrient-

limited grassland (56 of the 59 observations). Thus, it remains unclear whether vascular plant

diversity in highly productive, nutrient-rich wetlands may affect ecosystem functioning

positively, or even by the same mechanisms operating in grasslands (Engelhardt and Ritchie

2001).

One property that may set wetlands apart in their relation between diversity and functions

is the common local domination of a single – often exotic – vascular plant species. Indeed,

wetlands are particularly vulnerable to invasion by highly productive exotic species that form

near monocultures (Zedler and Kercher 2004). The negative impact of invasive macrophytes

on wetland diversity is undisputable, but it is unclear whether it is accompanied by a decrease

in ecosystem processes closely related to wetland ecosystem services. For example, nitrogen

retention may be comparatively greater in wetlands dominated by the invasive Phragmites

australis than in those that are more diverse (Hersher and Havens 2008). Also, large floating

beds of the invasive Trapa natans in the tidal Hudson River have been identified as

denitrification hotspots, removing significantly more nitrogen from the river than native

vegetation (Tall et al. 2011).

There are good theoretical reasons or indirect signs suggesting that increased plant richness

would result in increasing pollutant removal efficiency in constructed wetlands: better root

partitioning, complementary nutrient uses, increased bacterial diversity and activity, etc. Yet,

there is still little supporting empirical evidence for the positive effect of plant diversity on

pollutant removal. There are a few studies comparing removal in monocultures and

polycultures in experimental constructed wetlands, and their overall results are inconclusive

or inconsistent. Moreover, plant composition in wetland polycultures is difficult to maintain

due to community dynamics and the progressive dominance of the most competitive species.

Not surprisingly, macrophyte species selection in treatment wetlands is still mostly based on


established practices, assumptions and circumstantial evidence, and constructed wetlands are

still mostly planted with a single species.

Besides its potential effect on efficiency, plant diversity may have several other benefits

such as increased resilience to perturbation or diseases, esthetical value, better habitat, etc.

Yet, the relation between diversity and wetland effect on water quality remains an open

question that provides promising research avenues.

REFERENCES

Cardinale BJ et al. (2011) The functional role of producer diversity in ecosystems. Am J Bot 98: 572-

592.

Engelhardt, KAM. & Ritchie ME (2001). Effects of macrophyte species richness on wetland

ecosystem functioning and services. Nature 411:687–689.

Hershner C. & Havens KJ (2008) Managing invasive aquatic plants in a changing system: strategic

consideration of ecosystem services. Conserv Biol 22:544–550.

Tall L et al. (2011) Denitrification hot spots: dominant role of invasive macrophyte Trapa natans in

removing nitrogen from a tidal river. Ecol Applic 21: 3104-3114.

Zedler JB, Kercher S (2004) Causes and consequences of invasive plants in wetlands: Opportunities,

opportunists, and outcomes. Crit Rev Plant Sc 23:431-452.


Wastewater, examples on new organic contaminants, upcoming

metal(loid)s, nanomaterials & the transfer/transformation in

wetlands

Lars Duestera, Bjoern Meermann

a, Anne-Lena Fabricius

a, Michael Schluesener

a

and Thomas A. Ternesa

aFederal Institute of Hydrology, Department G2 - Aquatic Chemistry, Am Mainzer Tor 1,

56068 Koblenz, Germany

KEYNOTE

The presentation will provide a broad overview on metal and metalloid emerging

pollutants1 which are used in “new” industrial applications as well as on organic

compounds/particles that are under discussion to induce adverse environmental effects in

close to nature and constructed wetlands. The release scenarios and factors that impact the

environmental fate and transformation of potential pollutants will be addressed within the

presentation (figure. 1).

Figure 1: The four central themes and connecting links of the presentation.

I. “NEW” METAL(LOID)S

At a first glance and compared to the magnitude of man-made organic compounds that

show a potential to cause adverse environmental effects, inorganic compounds seem to have a

lower innovation potential in industrial application. This impression may change as soon as

one begins to think about speciation, fractionation2 and the availability of metals or metalloids

(metal(loid)s). Hence, for some metal(loid)s the occurrence of:

1. “new” metal(loid) organic species,

2. “new” fractions, e.g., nanoparticles,

3. or less commonly used metals in “new” industrial/medical applications,

is connected with certain concerns.

1Emerging pollutants: A substance currently not included in routine environmental monitoring programmes

and which may be candidate for future legislation due to its adverse effects and / or persistency

(http://www.norman-network.net/index_php.php?module=public/others/glossary#e_pollute).


2Chemical species: Chemical elements: specific form of an element defined as to isotopic composition,

electronic or oxidation state, and/or complex or molecular structure.

Fractionation: Process of classification of an analyte or a group of analytes from a certain sample according

to physical (e.g., size, solubility) or chemical (e.g., bonding, reactivity) properties (Templeton, Ariese et al.

2000).

As a simple example for point 1, zinc pyrithione is one of the most common biocides used

from personal care products to antifouling paints, but at the moment no analytical method is

available to allow a precise detection in different environmental matrices. Hence, it is

impossible to assess whether there is a risk toward wetlands, e.g., constructed wetlands in

water purification, or not. Point 2 is detailed in the next paragraph. Examples for point 3 are

the use of Gadolinium in MRT-contrast agents or the use of so far less used metals,

nowadays, applied in new technical application like semiconductors (e.g., thallium), micro

capacitators (e.g., niob) or in renewable energy applications (e.g., tellurium, germanium,

neodymium, table 1).

Table 1: Raw material emerging technologies (selected), modified after (EU Commission-Enterprise &

Industry 2010).

Element Application

Antimony micro capacitors

Cobalt Lithium-ion batteries, synthetic fuels

Gallium Thin layer photovoltaics

Germanium Fibre optic cable, IR optical technologies

Indium Displays, thin layer photovoltaics

Platinum Fuel cells, catalysts

Palladium Catalysts, seawater desalination

Niobium Micro capacitors, ferroalloys

Neodymium Permanent magnets, laser technology

Tantalum Micro capacitors, medical technology

For, e.g., technical critical elements, it becomes obvious that mining, the industrial

production (waste water) and recycling may pose risks to wetlands via the discharge of waste

waters. Especially low-tech recycling in emerging countries has to be considered in this

context.

II. ENGINEERED NANOMATERIALS

Changing from mostly speciation based scientific questions to fractionation based, within

the last years concerns on adverse environmental effects from unintentionally released

engineered nanomaterials (ENMs) were expressed by scientists (e.g., Wijnhoven, Peijnenburg

et al. 2009) and NGOs (etc-group 2010). An overview on nanomaterial definitions can be

found at JRC, 2010. A very pragmatic and handy definition on nanomaterials is that these

materials hold at least one dimension < 100 nm (e.g., a nano foil). This is sufficient to

understand the following considerations: Focusing on nanotubes (two dimensions < 100 nm)

as well as nanoparticles (three dimensions < 100 nm) and wetlands, waste waters from

industries/household and, hence, effluents from wastewater treatment plants (WTTPs) as well

as stormwater can be identified as potentially relevant sources of contamination (figure 2). As

most common nanomaterials in industrial applications and consumer products Ag and the

oxides of Ce, Fe, Si, Ti and Zn as well as carbon nano tubes (CNTs) were identified

(Piccinno, Gottschalk et al. 2012). First results on Ag (Kaegi, Voegelin et al. 2013; Hou, Li et

al. 2012), TiO2, SiO2 (Park, Kim et al. 2013) and ZnO (Hou, Xia et al. 2013) show separation

efficiencies > 90% from the waste water to the sludge during water treatment. This can be

taken as good news for wetlands, but variable process disturbances, like heavy rain events and

the release via stormwater, were not sufficiently addressed by now and leave space for


uncertainties. Beside challenges in environmental analytical chemistry and ongoing

discussions on the transformation of ENMs in WTTPs and in surface water environments, the

unknown input quantities from industries were recently identified as a general major

drawback in environmental risk assessments on ENMs (Hendren, Mesnard et al. 2012;

Piccinno, Gottschalk et al. 2012). With respect to the generally valid precautionary principle,

first results on the fate and effects of ENM in wetlands are now available (e.g., (Jacob,

Borchardt et al. 2013; Sharif, Westerhoff et al. 2013)).

Figure 2: Pathways and uncertainties in nanomaterial balances of WWTP. The picture shows the

Emscher WWTP in Germany (taken from COST Action ENTER ES1205

(http://www.cost.eu/domains_actions/essem/Actions/ES1205)).

III. MICRO AND MACRO PLASTICS

After and in parallel to a certain “nano–hype” in different scientific disciplines and also in

environmental sciences, the public and the scientific community were put on alert, by

environmental activists and by NGOs (e.g., http://plasticsoupfoundation.org/eng/beat-the-

micro-bead/), for a further group of particles – primary and secondary plastic particles. First

indicators and working areas was an increasing contamination of the oceans and shorelines

with plastics. In this context, primary particles are present in the size they were produced by

industries and secondary particles are products of weathering and fragmentation of bigger

plastic pieces (figure 3). With the rising awareness on environmental adverse effects from

plastics in the marine environment the amount of publications addressing this issue in marine

(which started ~ in the late 1980s (Liebezeit and Dubaish 2012) ) and, newly, freshwater

environments (Dubaish and Liebezeit 2013), wetlands (Cordeiro and Costa 2010) as well as

on quantitative and qualitative analyses of plastics in the environment, is increasing (e.g.,

(Claessens, Van Cauwenberghe et al. 2013; Hidalgo-Ruz, Gutow et al. 2012; Imhof, Schmid

et al. 2012)). As an example for a changing public perception and caused by public pressure

in the commission decision on criteria and methodological standards on good environmental

status of marine waters, in the descriptor 10, micro plastics are addressed (EU Commission

2010). In addition, at the beginning of 2013 Unilever agreed to phase out micro beads from

personal care products (The Guardian, 2013).


Figure 3: Examples for primary particles (left: polystyrene micro beads, mean diameter 230 µm in

vial) and secondary particles (right: plastic fragments (white spots) < 2 mm - 0.63 µm on a sieve in a freeze

dried river sediment, collected down stream a recycling facility in Germany). Micro beads are used in a

wide range of applications from lacquer to personal care products.

IV. EMERGING ORGANIC CONTAMINATANTS/MICROPOLLUTANTS

In contrast to the first three topics the scientific working area on organic

contaminants/micro pollutants and wetlands is dominated by questions on the removal

efficiency posed by wetlands by using them as a fourth purification stage in WTTP. Questions

on potential adverse effects are less often addressed. Depending on the intensity with which

this topic is already addressed in the conference, prior this lecture, the presentation will focus

on the degradation and transformation of organic analytes from personal care products and

pharmaceuticals as well as on biocides in constructed wetlands.

REFERENCES Claessens, M., L. Van Cauwenberghe, et al. (2013) "New techniques for the detection of microplastics

in sediments and field collected organisms." Marine Pollution Bulletin 70(1-2): 227-33.

Cordeiro, C. A. M. M. and T. M. Costa (2010) "Evaluation of solid residues removed from a

mangrove swamp in the Sao Vicente Estuary, SP, Brazil." Marine Pollution Bulletin 60(10): 1762-

1767.

Dubaish, F. and G. Liebezeit (2013) "Suspended Microplastics and Black Carbon Particles in the Jade

System, Southern North Sea." Water Air and Soil Pollution 224(2).

etc-group (2010). The Big Downturn? Nanogeopolitics. (http://www.etcgroup.org/fr/node/5245).

EU Commission Enterprise&Industry (2010). Critical raw materials for the EU

EU Commission (2010) (http://eur-

lex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:2010:232:0014:0024:EN:PDF)

Hendren, C. O., X. Mesnard, et al. (2012). "Estimating Production Data for Five Engineered

Nanomaterials As a Basis for Exposure Assessment." Environmental Science & Technology 45(7):

2562-2569.

Hidalgo-Ruz, V., L. Gutow, et al. (2012) "Microplastics in the Marine Environment: A Review of the

Methods Used for Identification and Quantification." Environmental Science & Technology 46(6):

3060-3075.

Hou, L. L., K. Y. Li, et al. (2012). "Removal of silver nanoparticles in simulated wastewater treatment

processes and its impact on COD and NH4 reduction." Chemosphere 87(3): 248-252.

Hou, L. L., J. Xia, et al. (2013). "Removal of ZnO nanoparticles in simulated wastewater treatment

processes and its effects on COD and NH4+-N reduction." Water Science and Technology 67(2):

254-260.

Imhof, H. K., J. Schmid, et al. (2012) "A novel, highly efficient method for the separation and

quantification of plastic particles in sediments of aquatic environments." Limnology and

Oceanography-Methods 10: 524-537.


Jacob, D. L., J. D. Borchardt, et al. (2013). "Uptake and translocation of Ti from nanoparticles in crops

and wetland plants." International Journal of Phytoremediation 15(2): 142-153.

JRC. (2010) Considerations on a Definition of Nanomaterial for Regulatory Purposes

(http://ec.europa.eu/dgs/jrc/downloads/jrc_reference_report_201007_nanomaterials.pdf).

Kaegi, R., A. Voegelin, et al. (2013). "Fate and Transformation of Silver Nanoparticles in Urban

Wastewater Systems." Water Research 47(12): 3866–3877.

Liebezeit, G. and F. Dubaish (2012). "Microplastics in Beaches of the East Frisian Islands Spiekeroog

and Kachelotplate." Bulletin of Environmental Contamination and Toxicology 89(1): 213-217.

Park, H. J., H. Y. Kim, et al. (2013). "Removal characteristics of engineered nanoparticles by activated

sludge." Chemosphere 92(5): 524-528.

Piccinno, F., F. Gottschalk, et al. (2012). "Industrial production quantities and uses of ten engineered

nanomaterials in Europe and the world." Journal of Nanoparticle Research 14(9).

Sharif, F., P. Westerhoff, et al. (2013). "Sorption of trace organics and engineered nanomaterials onto

wetland plant material." Environmental Science: Processes & Impacts 15(1): 267-274.

Templeton, D. M., F. Ariese, et al. (2000). "Guidelines for terms related to chemical speciation and

fractionation of elements. Definitions, structural aspects, and methodological approaches (IUPAC

Recommendations 2000)." Pure and Applied Chemistry 72(8): 1453-1470.

The Guardian, 2013; http://www.theguardian.com/environment/2013/jan/09/unilever-plastic-

microbeads-facial-scrubs?CMP=twt_gu)

Wijnhoven, S. W. P., W. Peijnenburg, et al. (2009). "Nano-silver - a review of available data and

knowledge gaps in human and environmental risk assessment." Nanotoxicology 3(2): 109-U78.


The Cartridge Theory: a Conceptual Approach to Horizontal-

Flow Wetlands’ Functioning

Roger Samsó and Joan García

GEMMA - Group of Environmental Engineering and Microbiology, Department of

Hydraulic, Maritime and Environmental Engineering, Universitat Politècnica de Catalunya-

BarcelonaTech, c/ Jordi Girona, 1-3, Building D1, E-08034, Barcelona, Spain.

([email protected] – [email protected])

INTRODUCTION

Numerical models are regarded by the scientific community as a potential tool to brighten

the black box to which Constructed Wetlands (CWs) have usually been assimilated. The

BIO_PORE model (Samsó and García, 2013a, b) is a numerical model for CWs resulting

from the combination of flow and transport equations and the biokinetic model Constructed

Wetland Model number 1 (CWM1) (Langergraber et al., 2009) within the COMSOL

MultiphysicsTM

platform. This model makes it possible to simulate bacterial growth and

pollutants degradation and transformations in wetlands and was developed with the aim of

improving the understanding of the internal functioning of horizontal subsurface flow CWs’.

In this paper we use simulation results obtained with the BIO_PORE model to develop

what we named “The Cartridge Theory” for horizontal subsurface flow constructed wetlands

(HSSFCWs), which is a high-level description of the functioning of these systems based on

the interaction between accumulated solids and bacterial populations.

Since the presented theory is mostly based on simulation results with BIO_PORE model,

we start by justifying the changes applied to the original formulation of CWM1 so that the

resulting growth of bacterial communities is consistent with existing population ecology

models. We do that by individually studying the evolution of the biomass of a single

functional bacterial group (fermenting bacteria) in a specific point near the inlet section of a

HSSFCW.

METHODS

Model description

BIO_PORE model is used to run all simulations. For details on model equations, main

hypothesis and assumptions, calibration and limitations, the reader is referred to Samsó and

(García, 2013a, b).

Pilot system

Simulations were run for a 10.3 m long and 5.3 m wide pilot wetland planted with

Phragmites australis. The granular medium consisted of fine granitic gravel (D60= 3.5 mm,

coefficient of uniformity= 1.7, initial porosity n= 40%) with a gravel depth of approximately

0.6 m at the inlet and 0.7 m at the outlet. The system was fed with urban wastewater

previously treated in an Imhoff tank.

Simulation strategy

Simulations represent the period comprised between start-up and the third year of

operation of the pilot system. Initial concentrations of all functional bacteria groups within the

wetland were set to 0.001mgCOD L-1

to represent start-up conditions. Constant values for

hydraulic loading rate (36.6 mm d-1

), water temperature (20 ºC) and influent pollutant

concentrations were used to facilitate interpretation of the model output. Influent


concentrations were extracted from data averages of an experimental study carried out in the

pilot wetland by García et al. (2005). The fractioning of the influent COD was made using

recommended values for primary effluents in ASMs (Henze et al., 2000).

RESULTS AND DISCUSSION

All results correspond to a point located in the inlet and near the water surface of the pilot.

Modeling bacterial growth in CWs

The concentration of fermenting bacteria (XFB, in mgCOD L-1

) through time at any

location using CWM1’s original formulation is obtained by solving Eq. 1:

(1)

The definition of the parameters, their values and units is given in Table 1.

Table 1. Values of the parameters of the equations describing the growth of fermenting bacteria.

Parameter Description Value Unit

Specific growth-rate 3 d

-1

bFB Rate constant for lysis 0.02 d-1

Initial concentration 0.001 mgCOD L-1

Saturation coefficient for SF 28 mgCOD L-1

Inhibition coefficient for SH2S 140 mgS L-1

Inhibition coefficient for SO 0.2 mgO2 L-1

Inhibition coefficient for SNO 0.5 mgN L-1

Saturation coefficient for SNH 0.01 mgN L-1

Therefore, with CWM1’s original formulation the growth of bacteria follows an

exponential tendency (Malthusian growth) and fermenting bacteria concentrations near the

inlet become unrealistically high after a very short simulation times (results not shown).

To prevent the formation of bacteria “hot spots” near the inlet section, we included a linear

function of the total bacterial density (Mbio) on the growth rate expression of each bacteria

group following the concept by Verhulst (1838) (Eq. 2):

(2)

Where, Mbio and Mbio_max (both in kgVS m-3

of granular material) are, respectively, the

sum of to the biomass concentration of all bacterial groups and the carrying capacity of the

environment. With this new expression, the growth of bacteria follows a logistic curve, and

bacterial concentrations stabilize once the carrying capacity of the system is reached (results

not shown).

However, wastewater and dead bacteria cells contain an inert fraction which is refractory

and accumulates in the granular media (causing its progressive clogging), which must

translate into decreasing bacteria concentrations with time. To simulate such phenomena, a

second negative feed-back function was added to Equation (1):


(3)

Where, MXIf (kgVS m-3

of granular material) and Mcap (kgVS m-3

of granular material) are,

respectively, the actual mass of inert solids and the maximum mass of solids that fit a cubic

meter of granular media. Fig. 1 shows fermenting bacteria concentrations over time resulting

from considering Mbio_max= 0.093 kgVS m-3

of granular material and Mcap= 6 kgVS m-3

of

granular media (both obtained from calibration in Samsó and García (2013a)).

Fig. 1. Fermenting bacteria concentrations (XFB) near the inlet section over time.

Simulation results on bacteria distribution and solids accumulation pattern in

HSSFCWs In this section bacterial and inert solids concentrations obtained with the previous

formulation are shown for the whole longitudinal section of the wetland. Fig. 2 shows that

bacteria communities are distributed in a rather narrow strip, occupying approximately a third

of the bed’s length at all times.

Fig. 2. Distribution of bacteria after 1 (top), 2 (middle) and 3 (bottom) years of operation.

On the other hand, Fig. 3 shows that inert solids coming from wastewater and from dead

bacteria cells initially accumulate near the inlet and progress towards the outlet with time.

Fig. 3. Distribution of accumulated inert solids after 1 (top), 2 (middle) and 3 (bottom) years

of operation.

kgCOD m-3

kgCOD m-3


THE CARTRIDGE THEORY

The Cartridge Theory states that a close interrelation exists between bacterial communities

and accumulated inert solids produced from bacteria dye-off and those contained in the

influent wastewater, which defines the most basic functioning patterns of all HSSFCWs. The

progressive accumulation of inert solids from inlet to outlet causes the displacement of the

active bacteria zone in the same direction (Fig. 2). This implies that wetlands have a limited

life-span which corresponds to the time when bacterial communities are pushed as much

towards the outlet that their biomass is not anymore sufficient to remove the desirable

proportion of the influent pollutants.

CONCLUSIONS

In this paper we presented a theory on the general functioning of HSSFCWs based on the

interaction between bacterial communities and accumulated solids (clogging) which was

derived from simulation results with BIO_PORE model.

The theory assimilates the granular media of HSSFCWs to a generic cartridge which is

consumed (clogged) with inert solids from inlet to outlet with time. The reduction of porosity

caused by the accumulation of solids causes the displacement of bacterial communities, which

are progressively pushed towards the outlet. According to this, the failure of a wetland occurs

when the active bacteria zone is located as close to the outlet section that its total biomass is

not sufficient to degrade an acceptable proportion of the influent pollutants.

This is the first time a high-level integrated description of the functioning of HSSFCWs is

made based on modeling results and represents an important step towards the complete

understanding of the functioning of these systems. This is also the first time the effect of

clogging by inert solids on bacterial communities is described.

ACKNOWLEDGEMENTS

This work was possible thanks to the funding from the Spanish Ministry of Innovation and

Science for the NEWWET2008 Project (CTM2008-06676-C05-01) and from the NAWATEC

FP7 Project (308336). Roger Samsó also acknowledges the scholarship provided by the

Universitat Politècnica de Catalunya (UPC).

REFERENCES García, J., Aguirre, P., Barragán, J., Mujeriego, R., Matamoros, V., Bayona, J.M., 2005. Effect of key

design parameters on the efficiency of horizontal subsurface flow constructed wetlands: long-term

performance pilot study. Ecological Engineering 25, 405-418.

Henze, M., Gujer, W., Mino, T., van Loosdrecht., M., 2000. Activated sludge models ASM1, ASM2,

ASM2D and ASM3. IWA Scientific and Technical Rep 9. London, UK: IWA Publishing.

Langergraber, G., Rousseau, D. P. L., García, J., Mena, J., 2009. CWM1: a general model to describe

biokinetic processes in subsurface flow constructed wetlands. Water Science and Technology 59 (9),

1687-1697.

Samsó, R., García, J., 2013b. Bacteria distribution and dynamics in constructed wetlands based on

modelling results. Science of the Total Environment. In press

Samsó, R., García, J., 2013a. BIO_PORE, a mathematical model to simulate biofilm growth and water

quality improvement in porous media: application and calibration for constructed wetlands.

Ecological Engineering 54, 116-127.

Verhulst, P.F., 1838. Notice sur la loi que la population suit dans son accroissement. Correspondance

Mathematique et Physique 10, 113–121.


The role and characterization of microbial communities in

wetlands for water pollution control

Kela Weber

Department of Chemistry and Chemical Engineering, Royal Military College of Canada,

Kingston (Ontario), K7K 7B4, Canada. ([email protected])

INTRODUCTION Microbial communities play an important role in wetlands designed for water pollution control

(Kadlec and Wallace, 2008; Truu et al., 2009; Faulwetter et al., 2009; Garcia et al., 2010). Microbial

communities 1) directly influence and contribute to contaminant removal, 2) develop biofilms which

can affect hydrological development, 3) have a close interaction with plant roots within the

rhizospheric region, and 4) can contribute to other beneficial or negative ancillary effects related to

treatment wetland operations. Treatment wetlands (TWs) house many different microenvironments

within a single system. Each microenvironment can have varying conditions, such as oxygen

concentration, redox potential, ionic strength, pH, nutrient availability, or pollutant concentration to

name a few. These variations allow for the development of diverse microbial communities within

different microenvironments of a treatment wetland. Figure 1 presents a simplified depiction of

microbial community interactions with plant roots, and the bed media.

Microbial communities can exist as free-floating microorganisms within the interstitial

spaces of the bed media or as anchored/attached colonies surrounding either the bed media or

integrated within the rhizosphere and root zone of the plants. It is generally accepted that the

interstitial microbial communities, although present, play a relatively small role in

contaminant removal when compared to rhizospheric or other biofilm bound microbial

communities. Depending on the oxygen concentrations and redox potential in a specific

region within a TW, different microbial communities will develop and therefore different

metabolic pathways will be responsible for the removal of pollutants from the water. For

Figure 1: Simplified depiction of microbial community interactions with bed media, plant roots, and organic wastewater components in a horizontal subsurface flow treatment wetland system. (Diagram not to scale)


example, microenvironments within the near-root zone (within 1mm of a root) of horizontal

subsurface flow wetlands can be largely aerobic (redox potential +250 to +700 mV), even

though the rest of the bed is dominated by anaerobic processes (redox potential +250 to -400

mV, Truu et al., 2009). The potential for localized conditions is one feature of TWs that has

allowed for unique and sometimes improved contaminant removal capabilities over more

conventional, high energy input, water treatment technologies.

Microbial communities play a large role in organic matter degradation in TWs (Faulwetter

et al., 2009). Both particulate organic matter (POM) and dissolved organic matter (DOM) are

removed within TWs through physiochemical and microbiological means. Preliminary

degradation steps are often initiated via exoenzymes excreted by microorganisms which help

cleave functional groups or specific bonds of large molecules, allowing for the eventual

internalization and metabolic utilization of the degradation products by microorganisms.

Microorganism metabolic products are ideally additional cellular mass, energy, CO2, and

water.

Nitrogen transformations and/or removal have long been a focus of TW design. The

general nitrogen removal process that is known to occur in many water treatment systems is

nitrification (aerobic) followed by denitrification (anaerobic). Hybrid systems such as

subsurface vertical flow (VF) followed by subsurface horizontal flow (HF) systems, or HF

followed by VF with a high recycle rate (Brix et al., 2006) have been aimed at accomplishing

significant total nitrogen removal. Some of the more recent advances in TW design such as

the fill and drain design (Nivala et al., 2013) have looked to take advantage of this idea by

varying saturation levels within single systems either temporally or spatially. Recent

discoveries have shown anaerobic ammonium oxidation (anammox), in addition to

heterotrophic nitrification coupled with aerobic denitrification to be of possible importance in

treatment wetlands as well. The unique conditions in TWs allows for both fast and slower

growing bacteria to develop which helps establish these distinctive processes (Wallace and

Austin, 2008).

Phosphorus is a required nutrient for microorganisms and therefore phosphorus is taken up

and stored within the cell mass of microbial communities. However, through cell death and

lysis this phosphorus will again be released at some point, making microbially mediated

phosphorus removal to be thought of as both limited and temporary (Garcia et al., 2010). On

the other hand, microbial communities in soil have been shown to assist in mineralising

organophosphate compounds (Truu et al., 2009). Thus, as a whole, the quantitative

contribution of microbial processes in the fate of phosphorus fate in TWs remains undecided.

Processes such as methanogenisis by methanogens, and sulphate reduction by sulphur

reducing bacteria (SRB) can lead to unwanted gaseous releases (CH4 and H2S) under

anaerobic or anoxic conditions. Finally, specific microbially mediated transformations have

been reported, such as BTEX and MTBE degradation, TCE dehalogenation, iron-oxidation,

and emerging contaminant transformation and/or mineralisation.

In addition to directly treating, utilizing, mineralizing or transforming pollutants in TWs,

microbial communities also play a role in terms of contaminant retention through the creation

of biofilms. The attachment or anchorage of microorganisms in TWs depends on the capsule

or slime layer surrounding the specific microbial communities developing in the TW, the

grain size of the bed media, the availability of roots or root hairs, and the local water velocity

in the immediate region. Microbial attachment/detachment occurs readily, with extracellular

polymeric substances (EPS) excreted into the slime layer or capsule region assisting

attachment, and shear stress working to detach the same microorganism. These EPS's are

made up largely of polysaccharides and proteins giving the microorganism an especially

sticky exterior. This sticky exterior also allows for the adsorption of contaminants from the

interstitial waters. This biofilm adsorption aids the physicochemical removal processes and


also provides non-motile microorganisms entrapped within biofilms access to a carbon and

energy source. Water velocity and the associated shear stress will have an effect on

microbiological development, but it may also affect the selection of specific groups or even

microbiological species developing within a system. This strictly biologically based biofilm

development has been documented in the literature and can have a significant effect on

overall system porosity (Weber and Legge, 2011). Porosity reduction based on

microbiological development also affects dispersivity (mixing) characteristics, and can lead to

preferential flow paths (short-circuiting), and even eventual clogging given specific

conditions (unpublished personal observations).

Microbial Community Characterization in Treatment Wetlands

The field of microbial community characterization has been through an immense growth

period within the last 30 years. Figure 2 summarizes the main categories of microbial

community characterization. Enumeration is one of the first characterization techniques

utilized in TWs. Originally this involved plate cultures and the subsequent counting of colony

forming units, filtering and dry weight measurements of total organic matter, and direct

counting and/or identification under a microscope (e.g. Petroff-Hauser counting). Later

developments included microbial staining techniques, flow cytometry, and eventually real-

time polymerase chain reaction (RT-PCR) (also known as quantitative PCR – qPCR).

Microbial activity methods were also developed and utilized very early in the field of TWs.

Although not always expressly described as microbial activity, activity measurements have

been used and described as soil respiration as far back as the 1980s. Respiration rates have

generally been measured in aerobic systems or using samples from aerobic regimes and have

most often tracked either O2 utilization rates, or CO2 production rates. Other activity

measurements include the direct or indirect quantification of adenosine triphosphate (ATP -

the main coenzyme used in cellular metabolism) or nicotinamide adenine dinucleotide

(NADH - coenzyme involved in cellular metabolism), and the quantification of extracellular

enzyme activities (eg. fluoresceine diacetate method).

Figure 3 summarizes, by methodology type, the number of peer reviewed publications

communicated each year which utilized a microbial characterization methodology. Some of

the first methods available for microbial community structure comparisons were fatty acid

methyl ester (FAME), and phospholipid-derived fatty acid (PLFA) analysis. Although not

used for direct identification of microorganisms they give the ability to compare or

differentiate complex microbial communities based on the specific make-up of the plasma

membrane surrounding microbial cells. A number of methods have been developed based on

the characterization of PCR amplified DNA segments from a complex microbial community.

Most methods utilize primers that amplify a highly conserved region of DNA encoding for the

16s ribosomal unit to gain an understanding of all prokaryotes in a sample; however other

regions or specific genes can be targeted to gain more specific information. Some of these

methods include denaturing gradient gel electrophoresis (DGGE), temperature gradient gel

electrophoresis (TGGE), and single-strand conformation polymorphism (SSCP), each of

which yield patterns of bands embedded within a gel which can then be excised and

Figure 2: Microbial community characterization techniques. (Cellular components not to scale)


sequenced. To gain a full understanding of microbial community structure sequencing is

required; however useful information regarding structural diversity can also be gained without

sequencing. Other methods that allow for community comparisons include terminal restriction

fragment length polymorphism (TRFLP), amplified rDNA (Ribosomal DNA) restriction

analysis, ribosomal intergenic spacer analysis (RISA), length heterogeneity PCR (LH-PCR),

and random amplification of polymorphic DNA (RAPD). Although all methods mentioned

can give useful information perhaps the most powerful method to be developed is

pyrosequencing which allows for simultaneous relative quantification and sequencing of all

targeted genes within a sample. Pyrosequencing holds great potential as it gives a complete

snap-shot of a sample’s microbial community structure in one simple method, but it is

currently the most costly method available, which in many cases can be prohibitive.

Clearly, microbial community structure can assist in gaining specific information regarding

the exact species or groups of microorganisms present in a system or sample. However, this

information can be difficult to relate back to implications or the exact quantitative role of the

microbial community present with regards to water pollution control or TW system

operations. In this regard microbial community function characterization is thought to be

more relatable to pollutant removal mechanisms and TW operations. Microbial community

function looks to gain an understanding of exactly what types and in what quantities the

microbial community is utilizing and excreting different compounds (see Figure 2 for a

pictorial depiction). It is through these basic functions that microbial communities interact

with different trophic levels and participate in different nutrient cycles in the environment,

and also offer pollutant removal capabilities in TWs. Rather than quantifying and identifying

DNA fragments within a sample, primers and probes can be developed for RNA segments.

Although RNA is more difficult to work with, it gives an actual indication of gene expression

and therefore an indication of a specific active function, rather than the potential for a specific

function when assessing DNA. qPCR and fluorescence in-situ hybridization have been used to

this end.

Community level physiological profiling (CLPP) is another functional characterization

method where the metabolic activity of a community sample is measured with relation to 31-

95 different carbon sources on a microtitre plate. With this method both a relative activity and

total metabolic potential for degrading a range of carbon sources is obtained.

The final functional approach is the use of microarrays such as the Geochip 3.0, to assess

the presence of anywhere from 20,000 to 60,000 genes via RNA (or DNA) segments using

specified probes on a small microscope slide. Although in its infancy this methodology also

holds great potential. With the expression of so many genes being assessed in a single sample,

full enzymatic pathways can begin to be assembled and assessed giving a more thorough

(although not directly measured) indication of overall function.

Figure 3: Summary of microbial community characterization publications in the field of treatment wetlands. Keywords: wetland, constructed wetland, treatment wetland, microbiology, microbial, microbiological (with all combinations). Databases: Compendex, Referex, Inspec, GEOBASE, GeoRef, Scifinder, Web of Science.


For many decades researchers and industry leaders alike have both understood and posed

many questions regarding the role of microbial communities in wetland systems. Future

research frontiers include both spatial and temporal analyses. At present we have many tools

available for microbial community characterization and the future holds many great

discoveries.

REFERENCES Brix, H., Arias, C.A., Johansen, N.H. (2006) Experiments in a two-stage constructed wetland system:

nitrification capacity and effects of recycling on nitrogen removal; in Wetlands-Nutrients, etals and

Mass Cycling. J. Vymazal (ed). Backhuys Publishers. Leiden, The Netherlands.

Faulwetter, J.L., Gagnon, V., Sundberg, C., Chazarenc, F., Burr, M.D., Brisson, J., Campera, A.K.,

Stein, O. (2009) Microbial processes influencing performance of treatment wetlands: A review.

Ecological Engineering. 35: 987–1004.

Garcia, J., Rousseau, D.P.L., Morato, J., Lesage, E., Matamoros, V., Bayona, J.M. (2010)

Contaminant Removal Processes in Subsurface-Flow Constructed Wetlands: A Review. Critical

Reviews in Environmental Science and Technology. 40: 561–661.

Kadlec, R.H., Wallace, S. (2008) Treatment Wetlands (2nd Ed.). CRC Press. Taylor and Francis

Group. Boca Raton, FL, USA.

Truu, M., Juhanson, J., Truu, J. (2009) Microbial biomass, activity and community composition in

constructed wetlands. Science of the Total Environment. 407: 3958-3971.

Nivala, J.A. , Headley, T., Wallace, S.D., Bernhard, K., Brix, H., van Afferden, M., Müller., R. (2013)

Comparative analysis of constructed wetlands: Design and construction of the ecotechnology

research facility in Langenreichenbach, Germany. Ecological Engineering. (in press).

Wallace, S., Austin, D. (2008) Emerging models for nitrogen removal in treatment wetlands. Journal

of Environmental Health. 71:10-16.

Weber, K. P. & Legge, R. L. (2011) Dynamics in the bacterial community level physiological profiles

and hydrological characteristics of constructed wetland mesocosms during start-up. Ecological

Engineering. 37, 666–677.


INVITED SPEAKERS

PR. SYLVIE DE BLOIS

Wetlands and wetland biodiversity in a changing climate

Sylvie de Blois is a professor of ecology at McGill University,

Associate Director of the McGill School of Environment, and a member

of the Center for Biodiversity Science in Quebec, Canada. Her research

focuses on plant and landscape ecology and in particular on the impact

of climate change on plant diversity. She is a co-leader of the CC-BIO

project, a major research initiative on climate change and biodiversity in

Quebec. She also leads CC-PEQ, a research project aimed at predicting

risks of biological invasion in a climate change context. She recently co-

authored a book on the impacts of climate change on Quebec Biodiversity (Changements

climatiques et biodiversité du Québec : vers un nouveau patrimoine naturel, Presses de

l’Université du Québec).

DIRK ESSER

25 years of treating raw sewage with reed bed filters in France - a personal history of

lessons learned

Dirk Esser studied agricultural science in Germany, where, as a scholar

of Prof. Kickhut, he first discovered treatment wetlands in the mid-

eighties. He then did a postgraduate training in France in sanitary

engineering, which allowed him to work with the research team of

IRSTEA (then Cemagref) on what was later to become the “French

System” reed bed filter. In 1991, after having negotiated a license

agreement with IRSTEA, he created his own company, SINT, in order to

bring this technology from research into widespread practical application.

He has been a major actor of the development of reed bed filters in

France, from a niche market into mainstream technology. By 2007, almost 500 treatment

wetlands in operation had been designed by his company. Dirk then managed to retire from

the daily work of running a company and to use his experience as an expert consultant. He

was recently involved in the founding of Global Wetland Technology – the international

association of leading specialist wetland technology companies.

DR CHRIS TANNER

Wetlands to control diffuse pollution at catchment scale

Chris Tanner is a principal Scientist at the National Institute of Water

and Atmospheric Research in New Zealand. Chris has undertaken

research and consultancy for over 20 years on the use of wetland and

pond ecotechnologies for treatment of domestic, agricultural and

industrial wastewaters, and diffuse run-off from urban, industrial and

agricultural land-uses. Working in a small country like New Zealand, he

has had the opportunity (i.e. has been forced) to work on a wide range of

different wetland systems and a diverse range of applications. He has

authored over 60 journal articles and book chapters, guest edited 2

journal special issues on wetlands and made more than 180 conference and workshop

presentations.


Wetlands and wetland biodiversity in a changing climate

Sylvie de Blois

McGill University, Department of Plant Science and the McGill School of Environment, 21

111 Lakeshore, Ste-Anne-de-Bellevue, Québec, Canada, H9X 3V9

[email protected]

In 2013, daily average CO2 concentration in the atmosphere reached a record level of 400

parts per million (ppm), a 27% increase compared to 1958 levels. Such concentrations are

considered by some experts as already beyond the safe upper limits above which physical

feedback mechanisms could drive the climate system into unstable states and cascading

catastrophic events. As a consequence of changes in CO2 and other greenhouse gases (GHG),

warming rate has increased in the last decades with all 10 of the warmest years on record

having occurred since 1998. Warming trends are not equally distributed on Earth, but the

largest temperature increases in this century are predicted to occur at high latitudes in

wetland-rich landscapes. Physical evidence for global warming is already noticeable in

melting glaciers, rising sea levels, diminished snow cover in the Northern hemisphere,

melting permafrost, changes in precipitation level, and frequent extreme weather events.

These processes are affecting ecosystems around the globe. Some of the consequences of

these global changes on wetlands and wetland biodiversity are reviewed here.

For wetlands, a variety of outcomes can be expected depending on regional trends in

temperature and precipitation and how these will interact with edaphic and topographic

conditions and modify the hydrological and biogeochemical cycles. Shifts in the quantity and

quality of wetlands around the globe are expected. Melting permafrost, for instance, will

change drainage patterns and the characteristics and distribution of Northern or alpine

wetlands. Increased evaporation and drying trends in some regions of the world are pushing

wetlands to unsustainable states, while frequent extreme weather events and flooding have an

impact on coastal wetlands and their services. Other consequences of climate change on

wetlands can be indirect. With warming, agriculture is likely to shift up north in some parts of

the world, bringing with it nutrient enrichment and declining water quality with consequences

on wetland sinks. An assessment at the global scale of how wetlands will respond to climate

change in this century and how these responses will in turn alter the many functions and

services of wetlands is needed.

One of the challenges for wetland ecologists and environmental engineers in this century is

to predict the fate of wetlands and associated biota and to manage wetlands in a context of

rapid climate change and high uncertainty. This amounts first to translating climate

predictions from a variety of climate models and GHG emission scenarios into pattern and

process familiar to ecologists and environmental engineers. Assuming that climate is one of

the main drivers of species distribution at broad spatial scale, species distribution models

(SDM) relating species location on a map with current temperature and precipitation patterns

have been combined with future climate scenarios to predict the potential location of suitable

climatic conditions for a variety of species, but rarely explicitly targeting wetland species.

Here, the results of SDM targeting obligate wetland plant species from peatland, marsh, and

swamp habitats in eastern North America are presented. Presence/pseudo-absence data were

coupled with recent interpolated weather station data for model training and testing using

growing degree days, total annual precipitation, and water balance (i.e., difference between

total annual precipitation and evapotranspiration) as predictors. The SDM were constructed

using a combination of regression (GLM, GAM, MARS), classification (CTA), and machine

mailto:[email protected]


learning (GBM, RF) methods, and an approach for optimal selection of future climate

scenarios from global and regional climate models was developed. A single consensus

prediction was produced for each species for 2041-2070 and 2071-2100. At the scale of this

study, the selected climate predictors allowed an accurate mapping of current plant species

distribution. Models for all species yielded high scores for predictive accuracy with AUC

values ranging from 0.83 to 0.99, increasing confidence in future projections. For future

projections, the spatially explicit maps produced from the models displayed considerable

latitudinal shifts in climatically suitable habitat for the two time periods and for all species

considered. A similar approach targeting invasive wetland plant species also showed

increased invasion risks in the study area. Results provide spatially-explicit information

relevant to our understanding of how novel wetland communities may arise in a rapidly

warming climate.

Climate change alters ecosystem process and function, triggering powerful feedback

mechanisms. Changes in biogeochemical cycles are of concern, especially regarding CO2 and

methane (CH4) emissions. Wetlands represent a major global source of CH4, a much more

potent GHG than CO2. There is evidence that CH4 emissions from wetlands in the past have

been not only responsive to climate, but may have in turn altered climate trajectory. Much

needed estimates of global CH4 emission from wetlands in a climate change context are

constrained by the lack of good data on wetland geographical distribution regionally and

globally. Changes in plant and microbial species composition under novel climates may as

well have implications for GHG emissions, with species showing different physiological

responses when environmental conditions change. Long-term monitoring in natural wetlands

is essential to improve predictions about the effects of climate change on wetlands, but

wetland mesocosms can also serve as model systems to test specific hypotheses about climate

change and wetland functions and services, including GHG emission and water quality

improvement. The challenges will be to scale up findings from mesocosms to natural systems

and to scale down climate predictions from global to local scales.


25 Years of Treating Raw Sewage with Reed bed Filters in France

- a Personal History of Lessons Learned

Dirk ESSER, SINT, La Chapelle du Mont du Chat, 73370, FRANCE,

([email protected])

The particularity and originality of the so-called “French System” is that we treat raw,

unsettled wastewater in a first treatment stage, thus integrating the treatment of primary

sludge into the reed bed treatment system, and that we do this with quite a reduced total

surface area (2 m²/p.e. for a full treatment in two vertical stages in series), as compared to

most other vertical flow reed-bed systems. This alleviates the operator from the recurrent task

of sludge management and disposal, as a well mineralized (OM reduction of about 60 %) and

dry (DM > 30 % in summer) humus layer has to be removed every 8 to 12 years on highly

loaded systems, and even less frequently on lower loaded system. This also largely avoids

odor problems, as the water and sludge treatment are fully aerobic. As other constructed

wetlands, these systems can be well integrated into the landscape. I think all these factors

explain the success of this system. I have been directly or indirectly involved in the design of

about 700 systems, and in 2012, according to the data base of the water-agencies, there were

around 2300 municipal reed bed filters operating in France (Lesavre, 2012), almost all of

them treating raw sewage. Most of them are “classical” two stage vertical flow reed bed

filters, but reed bed filters have also been used as a first treatment stage upstream of (mostly

already existing) pond systems, horizontal flow reed bed filters, rapid infiltration system (sand

filters), other soil-based treatment systems and sometimes trickling filters. Construction

companies such as EPUR NATURE and others have developed their own designs: single

stage systems, with or without recirculation, and have integrated saturated zones into their

design in order to enhance denitrification. Active filter materials (steel s slag, rock

phosphate) have been tested and start to be used for P-retention.

The “French system” also has been exported (I have been involved in projects in Spain,

Portugal, Switzerland, Belgium and Germany, as well as in the French overseas departments

and territories) and copied in other countries. More detailed descriptions of the “French

System” and its treatment performance are given in Molle et al. (2005) and, more recently, in

Troesch and Esser (2012).

When I joined them in 1988, my colleagues form the French national research institute

IRSTEA (then Cemagref) Alain Liénard and Catherin Boutin had already worked for some

years on reed bed filters. They had, somewhat by coincidence, started in the early 80’s to

monitor two plants which had been designed as “Max Planck Institute Systems” by the

German researcher Käthe Seidel. These systems where based on two vertical flow stages ins

series planted with reeds, both with several beds in parallel, the first one fed with raw sewage

according to the precepts of Mme Seidel, followed by a series of three horizontal filters,

planted with Scirpus and Iris. They worked rather well, but the first stage was prone to

clogging because of a thin sand layer on the top – in order to spread the water on surface of

the filter, as there was no batch feeding - and for this reason most designers and researchers

using or working with this system elsewhere ended up in installing a primary treatment

upstream (Burka and Lawrence, 1990)… or suppressed the vertical stages all together. In the

mid-eighties, the researchers from IRSTEA build their own pilot plant in Pont-Remy

(Lienard, 1987), still very close to the original Seidel design, but replaced the sand in the

vertical stages by fine gravel, and improved the treatment efficiency by introducing batch

feeding with pumps, which were necessary anyway due to the topography. This plant

showed, on the positive side, that a first stage vertical flow reed bed filter with a fine gravel

media would not clog when fed with raw sewage, if operated with alternating feeding and


resting cycles, and would not only filter very efficiently almost all incoming solids from the

wastewater, but also biologically degrade a big part of the dissolved pollution (organics and to

a lesser degree ammonium). But it also showed that a second stage vertical filter with only 5

cm of sand on a gravel media and without batch-feeding was not very effective. The

horizontal filters did not prove very efficient and clogged rather quickly, quickly in Pont-

Rémy as well as on both plants in Saint-Bohaire, even more so as they were long and not very

large, resulting in an unfavorable ratio between hydraulic section and gradient. In 1987, a full

scale plant was designed by IRSTEA and build in Gensac-la-Pallue, in Charentes, where 8

vertical flow reed beds, with a total surface area of about 1m² per p.e., were build in parallel

in order to provide a first treatment step for existing overloaded waste stabilization ponds.

This plant was monitored for several years by IRSTEA (Lienard et al. 1993) and is still in

operation today (Liénard 2010).

This was basically the state of the art of reed bed technology in France when I founded

SINT in 1991. We also had gathered some experience from the monitoring of rapid

infiltration (sand filter) systems (Guilloteau et al. 1993), which were becoming very popular

in France in the 90’s – but have almost completely disappeared today due to clogging

problems – less than 200 municipal systems in operation in 2012, according to the water

agencies data base (Lesavre, 2012).

We also learned from colleagues in other countries: in the late 80’s IRSTEA exchanged

with researchers and scientists from other countries, mainly from the UK and Denmark,

through the EC/EWPA Emergent Hydrophyte Treatment Systems Expert Contact Group

which drafted the first European design and operations guidelines for reed bed treatment

systems in 1990 (Cooper 1990). Personally, in the nineties, I also regularly attended informal

bi-annual meetings of a German-speaking group of experienced designers and researchers

"Erfahrungsaustausch Pflanzenkläranlagen" and the beginning cycle of regular IWA

conferences on Wetland Systems for Water Pollution Control (ICWS) also allowed for formal

and informal exchange of insights and knowledge. Finally, I also learned in the nineties when

working as part of an international design team for an EU founded pilot wetland in Greece

(European Commission, 2001). Many of the colleagues have retired, many are still working

with wetlands and some have become friends.

But when I started as a designer in 1991, a lot of questions still had to be addressed in

order to optimize our systems:

1. How should the beds be fed ? There was some batch-feeding of the first stages of

Gensac and Pont-Remy through pumps, but there was no siphon on the market to

work with raw waste water. Also feeding with a central gutter, as in these plants, did

not result in optimal distribution and these gutters needed regular cleaning.

Satisfactory solutions were rapidly developed and design values pragmatically

established for the batch volume, the number of feeding points and the batch feeding

velocity and were validated through monitoring and, for the batch volume, later

through labscale research (Molle, 2003)

2. How deep should the first stage beds be? They were rather shallow in Pont-Remy

and Gensac. Today we know that 30 cm of active filter with fine gravel seems to be

a minimum, and that most treatment takes place on the filter surface – alone about

half of the BOD and COD is withheld mechanically by surface filtration and the

biology is the most active in the surface layer and the accumulated sludge layer

(Chazarenc 2003, Chazarenc and Merlin, 2005). But the question how much

treatment performance improves with depth has still not been fully answered –

Molle and al. (2008) compared 60 and 80 cm deep filter media in a full scale

experimental arrangement in Evieu, but did not find any significant differences.

However, the work of Jaime Nivala and Tom Headley at UFZ with gravel based


filter material (Headley 2011), although with settled effluent, show that treatment

does improve with depth. My own theory is that there is a gain of treatment

efficiency with depth, but that this is not linear and that the marginal increase is

constantly decreasing, to become insignificant at one point. The effect of adding

aeration pipes in the filter media has not yet been clearly quantified either.

3. How should we design a second stage reed bed? After, at that time, an unsuccessful

attempt to improve the efficiency of the first stage up to full nitrification without

clogging and therefore build a horizontal filter for denitrification as a second stage,

we adopted a second stage vertical flow reed filter with a sand media to enhance

nitrification, based on our experience with rapid infiltration systems : these second

stages, with several, later only two, beds in parallel, were quite different in design

from the German/Austrian design of single filters with a rather thick unsaturated

sand layer and also from vertical flow filters in the UK, which had also evolved

from the Max Planck Institute System and consisted of a rather thin, more or less

saturated sand layer several which covered a much thicker, unsaturated gravel media

in which probably most aerobic treatment takes place, the role of the sand being

primarily to spread the water and slow down infiltration. In order to prevent

clogging, some of the parallel beds are put to rest, so that the organic matter on the

top of the filter and the biofilm inside the filter can undergo aerobic mineralization.

In our systems, we also use resting and feeding cycles on the second stage, but work

with batch feeding and an unsaturated sand filter media to do the treatment. This

means that any ponding on the sand layer of the second stage must be prevented.

Based on scientific evidence that treatment in vertical flow sand mostly occurs in

the first 10 cm and that hardly anything happens below 15 cm (Guilloteau, 1992;

Kunst und Flasche, 1996) because of limited oxygen supply, we limited our first

designs to 15 cm of sand. However, in practice, such thin layers did not give very

satisfactory results and we have build most of our systems with a sand layer of 30 or

40 cm depth. Recent work, for example of Weedon (2012) suggests that there might

be a gain in having a deeper sand layer of up to 60 cm. Our experience also showed

that the quality of the sand is crucial if the system is not to fail and that in some

areas of France, and some parts of the world, it is difficult to find suitable sand. So

there has been and still is some research on how to replace this second stage vertical

flow planted sand filter, if needed. (for example Prost-Boucle and Molle, 2010)

In the 90’s, reed bed treatment plants had not yet been implemented on a larger scale in

France, and for this reason, research funds remained very scarce: We, my colleagues from

IRSTEA and SINT, only managed to look at the reed beds as “black boxes” and knowledge

and design in these years progressed only by trial and error on full scale systems. The fact that

any significant errors on real sites would cost a lot of money excluded any bold steps in

unknown directions – we rather tried to progress gradually from what we knew and tried to

achieve higher treatment standards when we were not legally obliged to achieve them.

IRSTEA then sometimes managed to do monitoring in order to evaluate these improvements

and publish on them (for example Boutin et al. 1997). In parallel to our research, a team

from Université de Savoie/ESIGEC (now Savoie POLYTECH) did their own research on

their own pilot plant. The situation changed in the end of the 90’s, when our first plants had

proved to perform reliably for over 5 years and interest in these systems rose, through

publications and conferences of my colleagues from IRSTEA, by mouth to mouth propaganda

amongst decision makers of local communities, and also among some water agencies. The

latter pushed researchers and private operators to get together to establish guidelines resuming

best practice - SINT and one other small company were finally the only private actors who

worked on these guidelines, together with researchers an d water agencies, and they were not


published before 2005. The Rhone-Mediterranée-Corse Water Agency funded further

research of IRSTEA to progress in nitrogen and phosphorous removal in these systems. In

collaboration with SINT, the University of Savoy (Gerard Merlin) obtained a grant from the

Rhône-Alpes Region for a PhD student. So the first PhD students, Pascal Molle from IRSTEA

and Florent Chazarenc from University of Savoie, began to look inside the “black boxes” of

our reed bed filters and tried to understand how they work. Others have followed since.

IRSTEA managed to intensify their research work, covering not only N and P removal, but

also the limits of hydraulic and organic loading (Molle et al 2006, Boutin et al 2010) and, at

present, tries to further improve understanding through modeling (Petitjean et al. 2012). A

new center of competence has been created at the Ecole des Mines de Nantes in 2007 with the

arrival of Florent, who had come back from a post-doc in Canada. At the same time, as the

market grew, specialized design and build companies such as EPUR NATURE developed,

became more professional, created and drew on their own data bases and managed to finance

their own research, in collaboration with the above research institutes and sometimes

international teams.

Today, research still focuses largely on N and P removal. For N removal, the limits for

nitrification and nitrification dynamics have been studied and different configurations of

aerobic and anoxic zones have been and are tested (Prigent 2011, 2012; Millot et al. 2013).

Modeling of the kinetics of denitrification, depending on carbon and nitrate concentrations, as

well as temperature and probably other parameters is in progress (Morvannou et al. 2013).

Phosphorous removal has centered on the use of active filter materials as natural rock

phosphate and steel slags (Harouiya et al 2011, Barca et al. 2012) Both have specific

drawbacks, as natural rock phosphate needs some transformation and is expensive, whereas

steel slag is not as efficient and can considerably increase the pH of the water.

Also, research on the adaption of reed bed filters to tropical climates is financed by the

French government, in order to provide affordable treatment solutions to the overseas

departments and territories (Esser et al, 2010), while some systems have been operated and

monitored under relatively cold climatic conditions. BOD, COD, SS removal is regularly

monitored in “French type” reed bed systems but is not subject to any specific research work

as it is more than satisfactory and little progress seems necessary and possible. The

accumulation and transformation of sludge on the filter bed surface also largely works as

predicted and as the number of older plants is increasing, more and more experience with

sludge removal is gained. Only in a very few cases, sludge removal became necessary earlier

than anticipated, for reasons which could not always be elucidated. There have been rather

few failures of systems, clogging occurs on some occasions on the first stage, and is quite

normal if the system is heavily loaded in the start up phase, before the reeds and the biological

equilibrium are well established. Clogging on the first stage always seems reversible and to

my knowledge there has never been the necessity to change or refurbish filter material in the

first stage. On very few occasions, sand had to be replaced on the second stage. Such

problems, related to the quality of the sand or operational problems, mostly occurred quite

quickly after the commission phase. I only know of one case were an older plant – but by far

not the oldest – needed refurbishment of sand after a long ponding period. At present, I

conduct a survey to evaluate how nitrification on the oldest plants has evolved with the aging

of the filter material, especially on the second stage

In the discussion following my presentation, I would like to confront this experience with

that of other colleagues who also have their practical experience in designing systems, but

also with the scientific work of the wetland research community.

http://www.researchgate.net/researcher/59058305_A_Petitjean/


ACKNOWLEDGEMENTS

I wish to thank all scientists, researchers, colleagues and friends who have accompanied

my work throughout these years, with a particular mention to Alain Liénard, without whose

professional and personal support SINT would not have succeeded. On the whole, the

international “treatment wetland community” has always been very supportive and I wish to

express my gratitude towards them here.

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Lesavre, J. (2012): „Lagunes associées à des plantations de macrophytes en mulitifilières“

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International Office for Water in Paris on 24 Mai 2012

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of a recent plant in France. Water Sci Technol 19(12), 373–375.

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France. Water Science and Technology, 28(10), 159-167.

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recent developments in France. Proceedings of the 12th IWA international conference on

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Efficient Removal of Total Nitrogen by Constructed Wetlands? WetPol, Nantes, France, October 13

- 17, 2013

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vertical flow and horizontal flow constructed wetlands: a full-scale experiment study. Ecological

Engineering, 34, 23-29. Morvannou, A., Forquet, N., Troesch, S., Molle, P. (2013). How modelling improves the design of French

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Petitjean, A. ; Forquet, N. Wanko, A., J Laurent, J. Molle, P. Mosé, R., Sadowski A. (2012):

Modelling aerobic biodegradation in vertical flow sand filters: impact of operational considerations

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adapté aux filtres plantés de roseaux. PhD Thesis, Ecole des Mines de Nantes, 212 p. + annexes

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http://www.researchgate.net/publication/221870280_Modelling_aerobic_biodegradation_in_vertical_flow_sand_filters_impact_of_operational_considerations_on_oxygen_transfer_and_bacterial_activity


Wetlands to control diffuse pollution at catchment scale

Chris C. Tannera

a National Institute of Water and Atmospheric Research (NIWA), PO Box 11-115, Hamilton 3251, New

Zealand.

Wetlands are being increasingly recognised as tools to attenuate diffuse contaminant loads

across catchments. Although our knowledge of wetland treatment processes and performance

has increased significantly for individual wetlands, our ability to scale this up across

catchments is less developed. This workshop will explore current knowledge, future

directions and research needs relating to catchment-scale application of natural and

constructed wetlands for control of diffuse pollution. The key issues to be discussed include:

How the geographical and policy context in which we work effects our view of the

role of wetlands in diffuse pollution management, and can sometimes lead to

confusion between researchers, and land and water managers as to their practicality

and cost-effectiveness.

How differences in climate, landscape and land-use, and the consequent types of

contaminants generated influence wetland effectiveness, optimal location and

relative area required to control diffuse loads.

Our current capability to predict the treatment performance of wetlands receiving

variable diffuse flows and to scale-up across catchments, and

Fruitful new approaches, tools and alliances to speed progress in the future.

abstracts oral presentation posters · comparing outflows of the individual treatments for 2013,...

Documents